Spatial and temporal variation in the hydrochemistry of marine prawn aquaculture ponds built in acid sulfate soils, , .

Sarah Groves

A thesis submitted in fulfilment of the requirements for the degree of Doctor of Philosophy

Geography Program The School of Biological, Earth and Environmental Sciences The University of New South Wales October 2008

DEDICATION

This thesis is dedicated to special people in my life:

My dad, Michael; my mum, Elizabeth; and my sister, Emma. They have always provided me with love, encouragement and support in all that I do and for that I thank and love them very much.

And to those who were once in my life, but are no longer here; Nanna (Anne) Groves, and my friends Sonja Huddle and Claire Dean. They inspired me to not take anything for granted, to appreciate the little things, and above all, enjoy life.

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ABSTRACT

Many brackish water aquaculture ventures in Australia and overseas have established ponds in coastal regions with acid sulfate soils (ASS). Acid sulphate soils are known to leach relatively high concentrations of metals, acid (metal and H+ ion) and sulfur, however very little is known about how these leached elements affect the water quality of aquaculture ponds.

The main objective of this thesis was to describe the hydrochemical processes controlling the water chemistry in the water column and sediment pore water in the studied aquaculture ponds over time and space.

Water samples providing the spatio-temporal data were collected from the ponds with the use of adapted sampling methods commonly used in the groundwater environment. A transect of five nested piesometers was installed in two prawn ponds at Pimpama, , Australia. Each piesometer nest contained a multilevel with eight outtakes, a mini – horizontal, and a slotted piesometer. Water samples were collected from each nested piesometer on a bi-monthly basis over the prawn-growing season. The unstable elements and water quality variables (pH, Eh, DO, EC, water temperature) were measured in the field. Stable elements were analysed in the laboratory using ICP-OES and ICP-MS. Soil samples were collected at the end of the season for elemental analysis. A number of key sediment/water interactions and processes such as precipitation/dissolution reactions, oxidation-reduction reactions, photosynthesis, adsorption and seawater buffering were identified as important controls on pond water conditions. This is the first study to provide detailed hydrochemcial analysis of the pond water over time and space and aided in identifying that even shallow water bodies can be chemically heterogeneous.

Analysis of the water and sediment highlighted the selection of metals that can be associated with ASS and that are mobilised from pond sediments under certain chemical conditions. In Pond 7 Al, As, Ni and Zn concentrations were generally higher at the beginning of the grow-out season. Variability of the metal concentration was observed between the water column (0 – 1500 mm)

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and the pore-water (0 - -1000 mm). The highest concentration of Al (1044 μg/L) and Zn (104 μg/L) were sampled in the water column (approximately 400 mm from the surface of the pond). The highest concentration of As (130 μg/L) and Ni (73 μg/L) were sampled in the pore water sediment (associated with ASS). Elevated Mn and Fe2+ concentrations were also associated with the sediment pore water. The highest concentrations of Mn and Fe2+ were 4717 μg/L and 5100 μg/L respectively. In Pond 10, Ni concentrations (167 μg/L) were the highest at the beginning of the grow-out season. However, As (97 μg/L), Al (234 μg/L) and Zn (308 μg/L) were most concentrated during the middle of the cycle. The highest mean concentrations of these elements are As (63 μg/L), Al (91 μg/L) and Zn (69 μg/L) which are each associated with the sediment-water interface. These metals are integral in degrading the pond water quality and lead to a loss of beneficial algal blooms, a reduction in pond water pH, poor growth rates and high mortality in shrimp. It is also possible that the dissolved ions and precipitated compounds that are leached from the ASS are discharged into the adjacent coastal estuary of .

With knowledge obtained from this PhD study, effective management and treatment systems can be developed and implemented to minimise the impact of these soils on the pond system and the water discharging into natural coastal ecosystem.

Key Words: hydrochemistry, acid sulfate soils, iron, heavy metals, trace elements, Kuruma prawns, pond water, estuary.

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ACKNOWLEDGEMENTS

Firstly I would like to thank my supervisor and friend Dr Jesmond Sammut for his support throughout all stages of this thesis.

I would also like to thank the Australian Council for International Agricultural Research (ACIAR) for financially supporting this thesis. Without their support, the project would not have been possible.

A big thank you goes to the guys at Tomei Pty Ltd: Eduardo Viso, Thorbjorn Lyster, Kevin Kaywan, Hiro Ito, Kris Curtis, Scott Walter and Eu Teoh for allowing me to install piesometers and regularly sampling their ponds. They also assisted me a great deal in organising boats and accommodating me when sampling. They always made me feel welcome when I visited the farm and I hope that the information contained in this thesis helps them to better understand the chemical aspects of their ponds and improve the management of their acid sulfate soil.

I am grateful to the following people who supported me both in the field and in the laboratory. In the field – Shane Schofield, Jerzy Jankowski, Kavita Gosavi, Bethany O’Shea, Rosalind Desmier. Thanks for your friendship, support and the laughs. The monotony of field work was decreased by you being there beside me. In the laboratory – Dorothy Yu, Irene Wainwright and Ervin Slansky, and computer guru; John Owen. You also provided support, friendship and guidance, as well as reducing my stress levels. Thanks so much.

Thanks to Mike Melville and his team; Ben MacDonald and Annabelle Keene for the use of their soil corer and soil sampling advice.

Thanks to the people from the following institutions that provided me with data and their expertise: Kath Conway from the Bureau of Meteorology QLD and Len Cranfield from the Queensland Department of Natural resources and mines.

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Shane, thank you for the strength, wisdom, guidance, encouragement, love and support. Without you, the journey would have been a lot tougher.

A big thank you to my comrades in arms, my fellow PhD’ers and uni mates Bethany O’Shea, John Wischusen, and Louise Mazzaroli. Thanks for the hours of intellectual conversation, friendship, support, hot chocolate “discussion breaks”, chocolate cakes, party pies and above all, the laughs. You guys kept me going. You helped me grow as a person and into a researcher, and made my PhD years some of the most enjoyable and fulfilling of my life.

Thanks to the people at CSIRO Marine and Atmospheric Research for supporting me while finishing off the thesis part time. Thanks also to my dear friends from CSIRO; Frank Coman, Simon Tabrett, Simon Irvin, Jean Doak and Jan Wakeling for the shoulder to cry on, the intellectual debate, the tea breaks and the laughs.

Finally I would like to acknowledge my buddies Andrea Averkiou, Susanne Kruusamagi and Michelle Jones for their ongoing support of my thesis. You guys have also helped me enjoy my life away from the thesis, and have always provided me with support, love and friendship.

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LIST OF ABBREVIATIONS

ASS – Acid Sulfate Soils

AASS – Actual Acid Sulfate Soils

PASS – Potential Acid Sulfate Soils

ANZECC Guidelines – Australian and New Zealand Environment and Conservation Council Guidelines

DO – Dissolved Oxygen

EC – Electrical Conductivity

Eh – Redox Potential pE – Relative Electron Activity

ICP – MS - Inductively Coupled Plasma - Multi Spectral

ICP – OES - Inductively Coupled Plasma - Optical Emission Spectrometry

P. japonicus - Penaeus japonicus

AHD – Australian Height Datum

BDL – Below Detection Limit

SI – Saturation Indices

TDS – Total Dissolved Solids

UNSW – University of New South Wales

USGS – United States Geological Survey

BEES – School of Biological, Earth and Environmental Sciences

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TABLE OF CONTENTS

1 INTRODUCTION...... 25 1.1 Introduction...... 25 1.2 The Research Problem ...... 26 1.3 The Study Setting and Location ...... 27 1.4 Woongoolba/Pimpama Region - Geology and Literature ...... 28 1.4.1 Research on Pimpama and Woongoolba region ...... 29 1.4.2 Research at Tomei ...... 30 1.5 Aims ...... 30 1.6 Hypotheses ...... 31 1.7 The Database ...... 31 1.8 Project Outline ...... 34 1.9 Summary ...... 34

2 ENVIRONMENTAL SETTING – TOMEI FARM, SE QUEENSLAND ...... 36 2.1 Introduction...... 36 2.2 Location and Physiography ...... 36 2.2.1 Evidence of ASS in the Ponds ...... 38 2.3 Climate...... 39 2.3.1 Rainfall...... 41 2.3.2 Temperature ...... 42 2.3.3 Wind Patterns...... 42 2.4 Geological Setting...... 43 2.4.1 Deposition of the unconsolidated sediments ...... 45 2.4.1.1 The Formation of the Logan-Pimpama Coastal Area 46 2.4.1.2 Pimpama River and tributaries in the area 48 2.5 Geomorphology...... 49 2.5.1 Coastal Environments ...... 49 2.6 Recent History: Anthropogenic effects ...... 50 2.6.1 Land use: Development of Aquaculture Land...... 50 2.7 Summary ...... 51

3 REVIEW: FOCUS ON ACID SULFATE SOILS ...... 52 3.1 Introduction...... 52 3.2 Definition of Acid Sulfate Soils (ASS) ...... 52 3.2.1 Depositional Environments...... 53 3.2.1.1 Pyrite Precipitation 54 3.2.1.2 Pyrite Morphology 57 3.2.1.3 Pyrite Oxidation 58 3.2.2 Sulfides and ASS ...... 63 3.2.2.1 Iron monosulfide Formation 64 3.2.3 ASS – The Distribution of ASS ...... 65

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3.2.3.1 Australian Examples and Continued Awareness 68 3.2.3.2 Overseas Examples 76 3.2.4 ASS and its impact on the aquatic environment ...... 77 3.2.4.1 Estuaries, mangroves and associated land development 77 3.2.4.2 Effects of ASS on rivers 79 3.2.5 ASS and Aquaculture ...... 80 3.2.6 Trace elements ...... 80 3.2.6.1 Trace elements in Sediment pore water 80 3.2.6.2 Arsenic and Acid Sulfate Soils 84 3.2.6.3 Trace elements in Biota 85 3.2.7 The cost of Acid sulfate soils to the community...... 86 3.2.8 Remediation of Acid Sulfate Soils...... 87 3.3 Summary ...... 89

4 REVIEW: WATER CHEMISTRY IN ASS EFFECTED PONDS ...... 91 4.1 Introduction...... 91 4.2 Water...... 91 4.3 Sampling water (in lake/ponds) ...... 91 4.3.1 Seawater Chemistry – Focus on Trace Metals...... 95 4.3.1.1 Aluminium 96 4.3.2 Stratification and the Sediment-Water Interface ...... 99 4.4 Summary ...... 100

5 REVIEW: AQUACULTURE - SPECIES, WATER QUALITY AND POND MANAGEMENT ...... 101 5.1 Introduction...... 101 5.2 Farming Regimes ...... 101 5.3 Prawn Characteristics...... 102 5.3.1 Prawn Species ...... 102 5.3.2 Prawn Stocking and Grow-out ...... 104 5.3.3 Prawn Feed Requirements...... 105 5.4 Prawn Habitat Preferences...... 106 5.4.1 Penaeus japonicus – benthic inhabitants ...... 106 5.4.2 Aerators ...... 108 5.4.3 Dissolved Oxygen ...... 110 5.4.4 Nutrients ...... 114 5.4.4.1 Nitrogen 114 5.4.4.2 Phosphorus 118 5.4.5 Water exchanges ...... 118 5.4.6 Microorganisms...... 119 5.4.6.1 Sulfur Reducing Bacteria 122 5.4.7 Phytoplankton ...... 124 5.4.8 Disease...... 126 5.4.9 Harvesting...... 126 5.4.10 Pond water temperature during grow-out and shipment ...... 127 5.5 Aquaculture practices...... 127

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5.5.1 Managing aquaculture ponds...... 127 5.5.1.1 Filters 128 5.5.1.2 Managing Heavy Metals 128 5.5.1.3 Erosion 129 5.5.1.4 Synthetic Pond Liners 129 5.5.1.5 Chemical Use 130 5.5.2 Managing the Discharge Environment ...... 132 5.5.2.1 Natural Treatment 132 5.5.2.2 Settlement Ponds 133 5.6 Impact of Weather on the Pond Environment ...... 133 5.7 Summary ...... 134

6 FARM MANAGEMENT/GENERAL NOTES FROM THE FIELD...... 135 6.1 Introduction...... 135 6.2 Tomei Pond Logistics ...... 135 6.3 Water Exchange/Water Level Fluctuations ...... 143 6.4 Aeration ...... 145 6.5 Data Collected by Farm Management ...... 146 6.5.1 Temperature ...... 147 6.5.2 Salinity ...... 148 6.5.3 pH...... 149 6.5.4 Dissolved Oxygen (DO)...... 150 6.5.5 Secchi depth ...... 151 6.6 Feeding regime ...... 152 6.7 Bio-accumulation ...... 153 6.8 Pond preparation...... 154 6.9 Disposal of solid waste...... 155 6.10 Disposal of waste water...... 155 6.11 Chemical manipulation of the ponds and associated uncertainty ...... 156 6.12 Summary ...... 156

7 FIELD AND LABORATORY TECHNIQUES ...... 158 7.1 Introduction...... 158 7.2 Preparation for water sampling ...... 158 7.2.1 Design and Installation of Piesometers ...... 160 7.2.1.1 Water Column Sampling Installation 160 7.2.1.2 Sand Layer Pore-Water Instillations 163 7.2.1.3 Pond Base Pore-Water Installations 164 7.2.1.4 Support Poles 165 7.2.1.5 Sample Tube Extensions 166 7.3 Water...... 167 7.3.1 Sample Collection ...... 167 7.3.2 Chemical Analysis ...... 168 7.3.2.1 General Variables 168 7.3.2.2 Sample Collection, Storage and Analysis 169 7.3.2.2.1 Filtering 169

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7.3.2.2.2 Cations 169 7.3.2.2.3 Anions 170 7.3.2.3 Other Unstable Elements 172 7.3.2.3.1 Trace Elements (ICP-MS) 173 7.4 Sediment...... 174 7.4.1 Sampling sediment...... 174 7.4.2 Sample Collection ...... 174 7.4.3 Soil Core Analysis ...... 176 7.5 Hydrochemical Data ...... 179 7.5.1 Computer Manipulation ...... 179 7.5.2 Quality Control / Quality Assurance ...... 180 7.5.2.1 Charge Balance Error 181 7.5.2.2 EC / TDS 182 7.5.3 Statistics ...... 183 7.6 Summary ...... 183

8 POND SEDIMENTS...... 184 8.1 Introduction...... 184 8.1.1 Relationship between pore water and the adjacent sediments ...... 184 8.1.2 Impact on aquaculture species ...... 185 8.2 Classification of Soil Type at Pimpama ...... 185 8.3 Clay Types...... 185 8.3.1 Soil composition and clay mineralogy ...... 186 8.3.2 Chemical composition, structure and attributes of key clay minerals...... 186 8.4 Environmental Effects – adaptation and chemical interaction ...... 187 8.4.1 Impact of pH...... 188 8.5 Depositional Environment...... 189 8.5.1 Relationship between depositional environment and metal mobility...... 190 8.5.2 Site Description – Soil Characteristics ...... 190 8.6 Sediment Cores ...... 191 8.6.1 Sediment Core Description...... 192 8.6.2 Results from XRD...... 192 8.6.3 Sand and clay fraction analysis ...... 193 8.6.4 Water Content...... 196 8.6.5 Organic matter in Sediment Cores...... 197 8.6.6 Sediment pH ...... 197 8.6.7 Carbon, Nitrogen and Sulfur in Sediment Cores...... 198 8.7 Key Points to retain when considering the water chemical data...... 198 8.8 Summary ...... 199

9 POND HYDROCHEMISTRY – PHYSICAL VARIABLES...... 201 9.1 Introduction...... 201 9.1.1 pH...... 201 9.1.1.1 Pond 7 202 9.1.1.2 Pond 10 206

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9.1.1.2.1 Comments relating to pond pH 209 9.1.2 Water Temperature ...... 210 9.1.2.1 Pond 7 210 9.1.2.2 Pond 10 212 9.1.3 Dissolved Oxygen ...... 212 9.1.3.1 Pond 7 213 9.1.3.2 Pond 10 214 9.1.4 Eh (Redox)...... 215 9.1.5 Electrical Conductivity (Salinity)...... 218 9.1.5.1 Pond 7 219 9.1.5.2 Pond 10 220 9.2 Summary ...... 220

10 POND HYDROCHEMISTRY - MAJOR IONS & NUTRIENTS...... 222 10.1 Introduction...... 222 10.2 Establishing a representative seawater composition...... 222 10.3 Total Dissolved Solids ...... 224 10.4 Saturation Indices ...... 225 10.5 Sodium (Na+) ...... 226 10.5.1 Observed Sodium concentrations and potential source/sink...... 227 10.5.2 Sodium - spatial distribution in the ponds...... 228 10.6 Calcium (Ca2+)...... 231 10.6.1 Observed Calcium concentrations and potential source/sink ...... 232 10.6.2 Calcium - spatial distribution in the ponds ...... 235 10.7 Potassium (K+) ...... 238 10.7.1 Observed Potassium concentrations and potential source/sink ...... 238 10.7.2 Potassium - spatial distribution in the ponds ...... 240 10.8 Magnesium...... 243 10.8.1 Observed Magnesium concentrations and potential source/sink...... 243 10.8.2 Magnesium - spatial distribution in the ponds...... 244 10.9 Chloride ...... 246 10.9.1 Data summary and key processes ...... 247 10.9.2 Chloride - spatial distribution in the ponds...... 248 10.10 Sulfur Species...... 251 10.10.1 Data summary and key processes ...... 251 10.10.2 Sulfur - spatial distribution in Pond 7...... 253 10.10.3 Sulfur - spatial distribution in Pond 10...... 255 10.11 The Carbonate System ...... 259 10.11.1 Dissolved Inorganic Carbon versus Total Alkalinity ...... 261 10.11.2 Carbon Speciation...... 263 10.12 Nutrients...... 264 10.12.1 Pond 7 ...... 265 10.12.1.1 Ammonium 266 10.12.1.2 Nitrate 267 10.12.1.3 Phosphorus 268

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10.12.2 Pond 10...... 269 10.12.2.1 Ammonium 270 10.12.2.2 Nitrate 271 10.12.2.3 Phosphorus 272 10.13 Chemical Characteristics of Water - Statistical Analysis ...... 273 10.13.1 Spearman Correlation Coefficients ...... 273 10.14 Summary ...... 274

11 POND HYDROCHEMISTRY - TRACE ELEMENTS ...... 276 11.1 Introduction...... 276 11.2 Abundant Trace Elements ...... 277 11.2.1 Aluminium ...... 277 11.2.1.1 Concentrations and Distribution in the ponds 277 11.2.1.2 Potential source minerals 279 11.2.2 Manganese ...... 283 11.2.2.1 Concentrations and Distribution in the ponds 283 11.2.2.2 Potential source minerals 286 11.2.3 Iron ...... 289 11.2.3.1 Concentrations and Distribution in the ponds 290 11.2.3.2 Potential source/sink minerals 292 11.2.3.3 Impact on the surrounding estuary 297 11.2.3.4 Impact on the aquaculture species 298 11.3 Transition Metals ...... 298 11.3.1 Vanadium...... 298 11.3.1.1 Concentrations and Distribution in the ponds 299 11.3.1.2 Potential source/sink minerals 300 11.3.2 Chromium ...... 301 11.3.2.1 Concentrations and Distribution in the ponds 302 11.3.2.2 Potential source minerals 303 11.3.3 Cobalt ...... 304 11.3.3.1 Concentrations and Distribution in the ponds 304 11.3.3.2 Potential source minerals 307 11.3.4 Nickel...... 307 11.3.4.1 Concentrations and Distribution in the ponds 307 11.3.4.2 Potential source minerals 309 11.3.5 Molybdenum...... 309 11.3.5.1 Concentrations and Distribution in the ponds 309 11.3.5.2 Potential source minerals 310 11.4 Alkaline Earth Metals ...... 311 11.4.1 Strontium...... 311 11.4.1.1 Concentrations and Distribution in the ponds 311 11.4.1.2 Potential source minerals 312 11.4.2 Barium ...... 314 11.4.2.1 Concentrations and Distribution in the ponds 314 11.4.2.2 Potential source minerals 316 11.5 Other metallic elements...... 316 11.5.1 Copper...... 316 11.5.1.1 Concentrations and Distribution in the ponds 316 11.5.1.2 Potential source minerals 318

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11.5.2 Silver...... 318 11.5.2.1 Concentrations and Distribution in the ponds 318 11.5.2.2 Potential source minerals 320 11.5.3 Zinc...... 322 11.5.3.1 Concentrations and Distribution in the ponds 322 11.5.3.2 Potential source/sink minerals 323 11.6 Metalloids ...... 323 11.6.1 Arsenic...... 324 11.6.1.1 Concentrations and Distribution in the ponds 324 11.6.1.2 Potential source minerals 325 11.6.2 Boron ...... 326 11.6.2.1 Concentrations and Distribution in the ponds 326 11.6.2.2 Potential source minerals 327 11.7 Radioactive elements...... 327 11.7.1 Uranium ...... 327 11.7.1.1 Concentrations and Distribution in the ponds 328 11.7.1.2 Potential source minerals 329 11.8 Summary ...... 329

12 HYDROGEOCHEMICAL PROCESSES ...... 331 12.1 Introduction...... 331 12.2 Processes active in the pond base and dyke walls...... 331 12.2.1 Mineral dissolution and precipitation ...... 331 12.2.1.1 Sodium mineral dissolution/precipitation 331 12.2.1.2 Calcium mineral dissolution/precipitation 332 12.2.1.3 Potassium mineral dissolution/precipitation 332 12.2.1.4 Magnesium mineral dissolution/precipitation 333 12.2.1.5 Iron mineral dissolution/precipitation 333 12.2.1.6 Aluminium mineral dissolution/precipitation 333 12.2.1.7 Manganese mineral dissolution/precipitation 333 12.2.2 Redox (reversible reduction-oxidation) reactions...... 334 12.2.2.1 Reduction Reactions 335 12.2.2.2 Oxidation Reactions 339 12.2.3 Ion-exchange ...... 340 12.2.4 Mobilisation of metals ...... 341 12.3 Process active at the sediment-water interface...... 343 12.3.1 Mixing and dilution...... 344 12.3.1.1 Mixing 344 12.3.1.2 Dilution 344 12.3.2 Chemical buffering by seawater...... 345 12.3.3 Adsorption/Desorption reactions...... 345 12.3.4 Acidity /Alkalinity buffering ...... 346 12.3.5 Photosynthesis and respiration...... 346 12.4 Processes specific to the water column...... 347 12.4.1 Evaporation...... 347 12.4.2 Thermal stratification ...... 347 12.5 Summary of the differences between the chemistry of the two ponds ...... 347 12.6 Potential heavy metal affects on aquaculture species ...... 348

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13 CONCLUSIONS...... 350 13.1 Farm Management...... 352 13.1.1 Farm Practices that should be avoided in ASS...... 354 13.1.2 Actions to avoid problems in aquaculture...... 354 13.2 Recommendations for further research ...... 354 13.3 Final Comment...... 355

14 REFERENCES ...... 357

15 APPENDICES...... 389

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TABLE OF FIGURES

Figure 1: Location Map for the Woongoolba coastal flood plain and the Tomei Farm (marked in red)...... 28 Figure 2: Sample location map for the reference seawater samples taken from the Moreton Bay - Jumpinpin Bar Estuary System ...... 33 Figure 3: Woongoolba region showing Tomei Farm Location and adjacent waterways...... 36 Figure 4: Historical data for Ponds 7 and 10, (a) Prawn Production 1998-2000, and (b) averaged Total Actual Acidity (TAA) and Total Potential Acidity (TPA) from 2001 (modified from Gosavi, 2004)...... 38 Figure 5: Location of the weather station – 1) Logan City Water Treatment Station, 2) Gold Coast Seaway and 3) Rocky Point Sugar Mill (modified from Queensland road map)...... 40 Figure 6: Rainfall and temperature data summary for the three weather stations around the Tomei farm area (Gold Coast Seaway station #40764 and Logan WTS station #40854) ...... 41 Figure 7: Morning and afternoon wind speed and solar radiation data summary (Gold Coast Seaway station #40764 and Logan WTS station #40854) ...... 43 Figure 8: Tectonic Elements of the New England Orogen in Queensland: the study area is located on the most easterly tectonic subdivision - the Beenleigh Block (modified from Lohe, 1980) ...... 45 Figure 9: Relative sea level curve for Australia interpreted from the Papua New Guinea shoreline record: 150,000 years to present day (modified from Graham and Larsen, 1999)...... 46 Figure 10: Schematic cross section of the southern Moreton Bay Coastal Plain showing generalised lithologies and depositional environments (Tomei region is located near Jacob’s well) after Lockhart et al., 1998...... 48 Figure 11: Worm burrows in ASS facilitate the oxidation of sediments (modified from Kristensen, 2000) ...... 59 Figure 12: Global distribution of ASS (modified from www.cimmyt.org) ...... 66 Figure 13: Subset of the Queensland Government (Natural Resources and Mines) Acid Sulfate Soils Map for the Logan – Coomera region ...... 69 Figure 14: Queensland ASS distribution map ...... 70 Figure 15: Grab sampling methods undertaken in rivers and lakes (from Bartram and Balance, 1996)...... 92 Figure 16: Peeper (dialysis) in-situ water sampler...... 93 Figure 17: Conceptual model for anaerobic decomposition (biologically mediated reactions) of organic matter, showing the exchanges between the atmosphere, water column and sediments (modified from Jorgensen, 1983)...... 100 Figure 18: Sequence of electron acceptors...... 122 Figure 19: Volume and frequency of water in Ponds 7 and 10 ...... 144 Figure 20: Location of aerators, paddlewheels, water samplers and jetties in (a) Pond 7 and, (b) Pond 10...... 146 Figure 21: Daily morning and afternoon water temperature data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10...... 147 Figure 22: Daily morning and afternoon salinity data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10...... 148 Figure 23: Daily morning and afternoon pH data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10...... 150 Figure 24: Daily morning and afternoon DO data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10...... 151

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Figure 25: Daily Secchi depth data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10...... 152 Figure 26: Schematic of constructed water column piesometer showing bolt at base ...... 162 Figure 27: Schematic of sand and ASS pore water samplers...... 163 Figure 28: Schematic of pond showing the sediment sample collection points for Pond 7 and 10...... 175 Figure 29: Charge balance error plotted against Total Dissolved Solids (TDS) for (a) Pond 7 and; (b) Pond 10...... 182 Figure 30: Relationship between TDS and EC for (a) Pond 7 and; (b) Pond 10...... 183 Figure 31: Sediment core data summary for Pond 7 and Pond 10 (sand/clay ratios, moisture content, loss on ignition, results from the LECO analysis, and laboratory based sediment pH data) ...... 196 Figure 32: pH versus Depth from the Pond 7 water column and pore water samples over the three sampling periods...... 202 Figure 33: TDS versus pH - yellow shading shows the optimal pH range for the P. japonicus; (a) Pond 7 and (b) Pond 10 – note that the lowest pH values are in the ASS pore water and sand layer for both ponds and that the more acidic samples are typically more saline ...... 203 Figure 34: Time series of contoured pH transects for Pond 7 with 0cm representing the sand-water interface...... 205 Figure 35: Depth versus pH for the Pond 10 water column and pore water samples over four sampling periods ...... 207 Figure 36: Time series of contoured pH transects for Pond 10 with 0cm representing the sand-water interface...... 209 Figure 37: TDS versus Temperature with the P. japonicus comfort zone shown by the yellow shading for (a) Pond 7 and (b) Pond 10 – note the relationship between higher temperatures and higher TDS...... 211 Figure 38: Pond 7 water temperature (oC) over the sampling period showing a 7- 12oC temperature variation in the pond depending on which sampling round is considered...... 211 Figure 39: Pond 10 water temperature (oC) over the sampling period showing a 7- 15oC temperature variation in the pond depending on which sampling round is considered...... 212 Figure 40: Dissolved oxygen in (a) Pond 7 and (b) Pond 10 over time ...... 213 Figure 41: TDS versus DO for (a) Pond 7 and (b) Pond 10 (yellow shading is the zone of hypoxia for P. japonicus; note the same symbols are used in all cross plots, so refer to other graphs for a legend) ...... 213 Figure 42: TDS versus Eh for (a) Pond 7 and (b) Pond 10 – note that lower (more negative) measured Eh tends to accompany higher water salinity...... 215 Figure 43: pH versus Eh for (a) Pond 7 and (b) Pond 10 – note that lower measured Eh tends to accompany low (acidic) pH measurements ...... 216 Figure 44: Eh versus DO for (a) Pond 7 and (b) Pond 10 (see other bivariate plots for legend)...... 216 Figure 45: pE versus Eh diagram for pore water samples in Pond 10 ...... 217 Figure 46: Eh measurements from Pond 10 over the sampling period ...... 218 Figure 47: EC versus TDS showing the relationship between EC and salinity for the Tomei waters (a) Pond 7 and (b) Pond 10 ...... 219 Figure 48: EC versus Cl- for (a) Pond 7 and (b) Pond 10 (see other bivariate plots for legend)...... 220 Figure 49: Summary of CBE across the data set for (a) Pond 7 and (b) Pond 10 ...... 223 Figure 50: Calculated TDS versus Electrical Conductivity for (a) Pond 7 and, (b) Pond 10 ...... 224 Figure 51: Cl- versus TDS for (a) Pond 7 and, (b) Pond 10 ...... 225

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Figure 52: Na+ versus Cl- (meq/L) for (a) Pond 7 and, (b) Pond 10...... 227 Figure 53: Time series of contoured sodium concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02) ...... 229 Figure 54: Time series of contoured sodium concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)...... 231 Figure 55: TDS versus Ca2+ (mg/L) for (a) Pond 7 and (b) Pond 10 ...... 232 Figure 56: Cl versus Ca2+ (meq/L) for (a) Pond 7 and (b) Pond 10...... 233 Figure 57: Pond 7 pH versus Ca2+ (mg/L) for (a) Pond 7 and, (b) Pond 10...... 234 Figure 58: Time series of contoured calcium concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02) ...... 236 Figure 59: Time series of contoured calcium concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)...... 238 Figure 60: TDS versus K+ for (a) Pond 7 and (b) Pond 10...... 239 Figure 61: Cl- versus K+ for (a) Pond 7 and (b) Pond 10...... 239 Figure 62: Time series of contoured potassium concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02) ...... 241 Figure 63: Time series of contoured potassium concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)...... 242 Figure 64: TDS versus Mg2+ for (a) Pond 7 and (b) Pond 10...... 244 Figure 65: Time series of contoured magnesium concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02) ...... 245 Figure 66: Time series of contoured magnesium concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)...... 246 Figure 67: TDS versus Cl- for (a) Pond 7 and (b) Pond 10 ...... 247 Figure 68: Time series of contoured chloride concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02) ...... 249 Figure 69: Time series of contoured chloride concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)...... 250 2- 2- Figure 70: S versus SO4 (meq/L) for (a) Pond 7 and (b) Pond 10...... 252 Figure 71: TDS versus S2- (mg/L) for (a) Pond 7 and (b) Pond 10...... 252 2- Figure 72: TDS versus SO4 for (a) Pond 7 and (b) Pond 10...... 253 2- Figure 73: Time series of contoured SO4 concentration profile for each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02) ...... 254 Figure 74: Time series of contoured S2- concentration profile for each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02)...... 255 2- Figure 75: Time series of contoured SO4 concentration profile for each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and Jun-02)...... 256 Figure 76: Time series of contoured S2- concentration profile for each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and Jun-02)...... 258 2- Figure 77: SO4 (mg/L) versus pH for (a) Pond 7 and (b) Pond 10...... 258 Figure 78: S2- (mg/L) versus pH for (a) Pond 7 and (b) Pond 10...... 259 Figure 79: S2- (mg/L) versus DO for (a) Pond 7 and (b) Pond 10...... 259 Figure 80: Bjerrium plot (Drever, 1997)...... 260 Figure 81: Chloride, pH and Alkalinity verses DIC for Pond 7 (a, b, c) and Pond 10 (d, e, f) ...... 262

Figure 82: CO2 (mg/L) versus depth for (a) Pond 7 and (b) Pond 10 (note: the red dashed line represents the sediment-water interface) ...... 263 - Figure 83: HCO3 (mg/L) versus depth for (a) Pond 7 and (b) Pond 10 (note: the red dashed line represents the sediment-water interface) ...... 264 2- Figure 84: CO3 (mg/L) versus depth for (a) Pond 7 and (b) Pond 10 (note: the red dashed line represents the sediment-water interface) ...... 264

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+ + Figure 85: (a) pH versus NH4 (b) TDS versus NH4 for Pond 7 ...... 266 + + Figure 86: (a) Eh versus NH4 and (b) DO versus NH4 for Pond 7 ...... 267 - - Figure 87: (a) pH versus NO3 (b) TDS versus NO3 for Pond 7...... 267 - - Figure 88: (a) Eh versus NO3 (b) DO versus NO3 for Pond 7 ...... 268 3- 3- Figure 89: (a) pH versus PO4 (b) Eh versus PO4 for Pond 7...... 269 3- Figure 90: DO versus PO4 for Pond 7...... 269 + + Figure 91: (a) pH versus NH4 (b) TDS versus NH4 for Pond 10 ...... 270 + + Figure 92: (a) Eh versus NH4 (b) DO versus NH4 for Pond 10 ...... 271

Figure 93: TDS versus NO3 for Pond 10...... 272

Figure 94: (a) DO versus NO3 (b) Eh versus NO3 for Pond 10 ...... 272 3- Figure 95: (a) pH versus PO4 (b) Vivianite Calculated Saturation Indices for Pond 10 ...... 273 3- 3- Figure 96: (a) Eh versus PO4 (b) DO versus PO4 for Pond 10...... 273 Figure 97: (a) AlTOT versus TDS for Pond 7 (note: the red dotted line represents the ANZECC (2000) trigger value for Al TOT). All data to the right of the red line exceeds the guideline. The two pink seawater crosses to the right of the red dotted line are samples taken from the Tomei intake water and estuary. The two pink seawater crosses to the left of the line are reference seawater from the GBR and Hem. (b) AlTOT versus TDS for Pond 10 ...... 278 Figure 98: (a) pH versus AlTOT for Pond 7, (b) pH versus AlTOT for Pond 10...... 279 Figure 99: Saturation Indices for Aluminium bearing minerals in Pond 7 ...... 281 Figure 100: Saturation indices for Aluminium bearing minerals in Pond 10 ...... 282 Figure 101: MnTOT concentrations versus TDS for (a) Pond 7 (Note: the red dotted line shows the ANZECC Guidelines (2000) trigger concentration) and (b) Pond 10 ...... 284 Figure 102: pH versus MnTOT for (a) Pond 7 and (b) Pond 10...... 284 Figure 103: DO versus MnTOT for (a) Pond 7 and (b) Pond 10...... 285 Figure 104: Eh versus MnTOT for (a) Pond 7 and (b) Pond 10...... 285 Figure 105: (a) MnTOT versus Fe2+ for Pond 7 and (b) MnTOT versus Fe3+ Pond 7 ...... 286 Figure 106: Saturation Indices for manganese minerals in Pond 7...... 288 Figure 107: Saturation indices for Manganese minerals in Pond 10 ...... 289 Figure 108: pH versus FeTOT in (a) Pond 7 and (b) Pond 10 (use legend on (a) for both graphs)...... 291 Figure 109: TDS versus Fe2+ for (a) Pond 7 and (b) Pond 10...... 292 Figure 110: Saturation Indices for iron minerals in Pond 7 ...... 295 Figure 111: Saturation indices for iron minerals in Pond 10 ...... 296

Figure 112: Depth versus CO2 for (a) Pond 7 and (b) Pond 10...... 297 Figure 113: Total vanadium versus TDS (with ANZECC Guidelines trigger value marked on in a red dotted line) (a) for Pond 7, and (b) for Pond 10 ...... 300 Figure 114: Vanadium versus arsenic in (a) Pond 7 and (b) Pond 10...... 301 Figure 115: Vanadium compared to (a) Boron, (b) Chromium and (c) Lithium in Pond 7 ...... 301 Figure 116: Total chromium versus TDS (with ANZECC Guidelines trigger value marked on in a red dotted line) (a) Pond 7 and (b) Pond 10 ...... 303 Figure 117: Eh versus Total chromium (with pink dotted line to show the transition from anoxic to oxic waters) (a) Pond 7 and (b) Pond 10...... 304 Figure 118: TDS versus Total Cobalt for (a) Pond 7 and (b) Pond 10...... 305 Figure 119: Eh versus Total Cobalt for (a) Pond 7 and (b) Pond 10 ...... 306 Figure 120: pH versus Total Cobalt for (a) Pond 7 and (b) Pond 10 ...... 306 Figure 121: Strontium versus Total Cobalt for (a) Pond 7 and (b) Pond 10...... 307

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Figure 122: Total Nickel versus TDS for (a) Pond 7 and (b) Pond 10 ...... 308 Figure 123: pH versus Ni for (a) Pond 7 and (b) Pond 10...... 308 Figure 124: Mo versus pH for (a) Pond 7 and (b) Pond 10...... 310 Figure 125: Chloride versus Sr (mg/L) for (a) Pond 7 and (b) Pond 10...... 312 Figure 126: pH versus Sr for (a) Pond 7 and (b) Pond 10 ...... 312 Figure 127: DO versus Sr for (a) Pond 7 and (b) Pond 10...... 313 2- Figure 128: CO3 versus Sr for (a) Pond 7 and (b) Pond 10...... 313 2- Figure 129: SO4 versus Sr for (a) Pond 7 and (b) Pond 10...... 313 Figure 130: pH versus Total Ba in (a) Pond 7 and (b) Pond 10 ...... 315 Figure 131: Total Cu versus TDS in (a) Pond 7 and (b) Pond 10...... 317 Figure 132: Total Cu versus pH in (a) Pond 7 and (b) Pond 10 ...... 318 Figure 133: AgTOT versus TDS (with ANZECC Guidelines marked on) for (a) Pond 7 and (b) Pond 10 ...... 319 Figure 134: AgTOT versus pH for (a) Pond 7 and (b) Pond 10 ...... 320 Figure 135: Saturation Indices for Acanthite in Pond 7...... 321 Figure 136: Saturation Indices for silver minerals in Pond 10...... 322 Figure 137: ZnTOT versus TDS (with ANZECC Guidelines in red) (a) Pond 7 and (b) Pond 10 ...... 323 Figure 138: AsTOT versus TDS for (a) Pond 7 and (b) Pond 10...... 325 Figure 139: pH versus AsTOT for (a) Pond 7 and (b) Pond 10 ...... 325 Figure 140: TDS versus B for (a) Pond 7 and (b) Pond 10 (Blue dashed line 4.6 mg/L average seawater concentration from Hem (1989))...... 327 Figure 141: pH versus UTOT for (a) Pond 7 and (b) Pond 10 (blue dashed line shows the average seawater concentration of 3 μg/L (Hem, 1989)) ...... 328 Figure 142: Average trace metal concentrations in the ponds at Tomei. Notice that the dominant trace elements in the sediment pore water were Fe2+ and Fe3+...... 330

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TABLE OF PLATES

Plate 1: Air photo of the Tomei Farm showing layout of the aquaculture ponds, intake and discharge points and adjacent seawater sample locations (Natural Resources and Mines Qld, 2002)...... 37 Plate 2: Iron oxyhydroxide staining on pond walls (a) and (b): this is a common feature associated with the oxidation of ASS when exposed to the atmosphere...... 39 Plate 3: Scanning Electron Microscope (SEM) image of framboidal pyrite which is the common form associated with organic matter (from Preda and Cox, 2004); (a) fresh pyrite crystals, (b) Framboid affected by acidity over several months of exposure. Many crystal edges are etched and covered with iron phases...... 58 Plate 4: SEM images of Jarosite (a) synthetic Jarosite, (b) naturally occurring Jarosite crystal taken from www.tltc.ttu.edu and www.earthsciences.uq.edu.au...... 63 Plate 5: Scanning Electron Micrograph (SEM) of colloidal greigite (magnetic iron sulfide) taken from www.earth.leeds.ac.uk...... 65 Plate 6: Yellow arrow points at a fungus (Aphanomyces invadans) attacking epithelial fish cells (Sammut, 1998) ...... 67 Plate 7: Acid sulfate soil exposure resulting in the death of vegetation down slope (Hey, 1999) ...... 68 Plate 8: Corrosion of concrete pipes (Hey, 1999)...... 68 Plate 9: Marine sediments have a poor load bearing capacity (from Hey, 1999)...... 71 Plate 10: Drying out of marine sediments leads to shrinking and cracking resulting in a mosaic, prismatic pattern on the surface and extending deep in the soil profile (from Hey, 1999)...... 72 Plate 11: An indicative mineral of oxidised ASS is jarosite (from Hey, 1999) ...... 72 Plate 12: Major fish kill in Pimpama River in 1996 (from Hey, 1999)...... 75 Plate 13: (a) Lyngbya majuscula floating in Moreton Bay, (b) Proliferation of the species shown at low tide – this is associated with Fe leaching out of adjacent ASS ...... 76 Plate 14: Corrosion of concrete pylon in the Pimpama River (from Hey, 1999)...... 76 Plate 15: Penaeus japonicus is the main species cultured at the Tomei prawn farm...... 103 Plate 16: Aspirators at Tomei ...... 108 Plate 17: Paddlewheel aerators at Tomei ...... 109 Plate 18: Air photo from 1944 which predates the establishment of the Tomei prawn farm, the 2002 farm outline is shown in yellow. Note that in 1944 most of the area was vegetated by mangroves...... 136 Plate 19: Air photo from 1970 which predates the establishment of the Tomei prawn farm; in 1970, there is little change in the land use pattern...... 137 Plate 20: Air photo from 1990; the south eastern section was excavated in 1987 for the commencement of marine aquaculture activity, and the western section was being used for cane farming...... 137 Plate 21: Air photo from 1995; the southern ponds were being actively used for marine aquaculture, and the western section was still being used for cane farming...... 138 Plate 22: Air photo from 2002; the Tomei prawn farm was fully established and operational. This photograph was taken in February so coincides with the 2nd sampling round – note the distribution of paddle wheel and aspirators...... 139

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Plate 23: (a) Pond 10 showing the irregular distribution of vegetation (salt marsh grass and succulents) on the dyke wall, (b) close up of more densely vegetated succulent growth on one of the pond dyke walls ...... 140 Plate 24: Synthetic pond wall liner is used only on the dyke walls at the Tomei prawn farm ...... 141 Plate 25: Imported sand is spread on the pond base (approximately 10-20cm thickness) prior to filling in preparation for the new growout season at the Tomei prawn farm ...... 141 Plate 26: Benthic fauna and flora, (a) single Bursatella leachi; (b) proliferation of Bursatella leachi; (c) egg cluster of Bursatella leachi; (d) iron stained tube worm cast Serpula rubens; (e) common example of the benthic organisms found on the pond floor at the end of the season...... 154 Plate 27: Pond Preparation at Tomei...... 155 Plate 28: Construction of piesometers off site, (a) construction equipment; (b) Food grade plastic tubing (measured, cut and filters installed); (c) piesometer support poles in centre of picture with sediment sample elbows; (d) installed sampling point with filter sock installed; (e) completed piesometer (top in the background) with hammer as scale...... 159 Plate 29: Installation of water sampling devices...... 160 Plate 30: Plastic tubing taped to the outside of the conduit during piesometer construction...... 162 Plate 31: Stages of construction of the sand pore water mini-samplers...... 164 Plate 32: (a) cutting 40mm PVC pressure pipe into 3m lengths; (b) slotting the 40mm PVC intake section of the ASS clay pore water piesometer...... 165 Plate 33: Installation and use of wooden support poles during sampling in the ponds at Tomei...... 166 Plate 34: Tube roll containing plastic tubes which are attached to pond piesometer for sample collection on dyke wall...... 167 Plate 35: Chloride colorimetric titrations carried out in the laboratory ...... 172 Plate 36: Reduced ions were analysed in the field with the use of (a) HACH spectrophotometer and, (b) colorimetric methods ...... 173 Plate 37: (a) Stainless steel sediment sampler and plastic sample sleeve; (b) one way inserts/valves and stainless steel sample cutter; (c) complete unit with pipe wrench...... 176 Plate 38: Spear probe inserted into the core for pH readings ...... 178 Plate 39: Particle Size Analysis (PSA) equipment...... 178 Plate 40: Weighing out sediment fractions after PSA ...... 179 Plate 41: Black monosulfide contained at the base of a sediment core from Pond 10 is typical of those collected from Tomei. The sand layer at the top of the core is added during pre-season preparation and shows evidence of iron oxide staining...... 192

Plate 42: CaCO3 worm structures on the floor of the pond bottom and showing iron oxide staining...... 234

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TABLE OF TABLES

Table 1: Water sample database...... 32 Table 2: Summary of data used in this project ...... 33 Table 3: Oxyhydroxides, hydroxides and oxides, sulfides and sulfates from Fitzpatrick et al. (1993)...... 54 Table 4: Composition of anaerobic and aerobic pore water (after Maab et al., 1985 – in Calmano et al., 1990 pg 509)...... 84 Table 5: Information of species, geographic range and life history (adapted from Benzie, 2000)...... 104 Table 6: Mechanisms for increasing the concentration of DO in aquaculture ponds (adapted from Boyd, 1998)...... 112 Table 7: Concentration of oxygen and physical response from prawn...... 113 Table 8: Heterotrophic bacteria that fixate nitrogen in marine environments (modified from Herbert, 1999)...... 120 Table 9: Ponds 7 and 10 Dimensions ...... 140 Table 10: Summary Table for Tomei Prawn Farm 2001/2002 growing season...... 143 Table 11: Summary of field data collected during study ...... 158 Table 12: Numbering configuration for piesometers...... 161 Table 13: HACH DR/2000 Standard Methods used for unstable elements...... 173 Table 14: Software used during this study...... 179 Table 15: Descriptive statistics for the particle size analysis for samples from Pond 7 (n=24) and Pond 10 (n=14) ...... 193 Table 16: Range of recorded Physical Variables recorded for Pond 7...... 203 Table 17: Range of recorded Physical Variables in Pond 10...... 207 Table 18: Average Dissolved Oxygen (DO in mg/L) concentrations for Pond 10 ...... 215 Table 19: Mean concentrations of Ca2+ in the water column and pore water from Pond 7 ...... 237 Table 20: Descriptive statistics for Nutrients in Pond 7 (in mg/L) (n=105)...... 265 Table 21: Descriptive statistics for Nutrients in Pond 10 (in mg/L) (n=134) ...... 270 Table 22: Combined table of ANZECC Guidelines (2000) for maximum concentrations of elements and detection limits of the ICP-MS used in this study ...... 276 Table 23: Descriptive statistics for Fe2+, Fe3+ and total iron (Pond 7 n=105 and Pond 10 n=134)...... 291 Table 24: Iron concentrations from the seawater samples taken near Tomei (note: see Figure 2 for reference seawater sample location map) ...... 292 Table 25: Descriptive statistics for vanadium (Pond 7 n=105 and Pond 10 n=134)...... 299 Table 26: Descriptive statistics for Chromium in Pond 7 (Pond 7 n=105)...... 302 Table 27: Descriptive Statistics for Molybdenum in Pond 7 (n=105) ...... 309 Table 28: Descriptive Statistics for Strontium (Pond 7 n=105 and Pond 10 n=134) ...... 311 Table 29: Descriptive Statistics for Barium (Pond 7 n=105 and Pond 10 n=134) ...... 315 Table 30: Descriptive Statistics for Copper (Pond 7 n=105 and Pond 10 n=134)...... 317 Table 31: Descriptive Statistics of AgTOT (Pond 7 n=105 and Pond 10 n=134)...... 319 Table 32: Descriptive Statistics for Zinc (Pond 7 n=105 and Pond 10 n=134)...... 322 Table 33: Descriptive Statistics for Arsenic in Pond 7 (Pond 7 n=105 and Pond 10 n=134) ...... 324 Table 34: Descriptive Statistics for Boron (Pond 7 n=105 and Pond 10 n=134) ...... 326

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Table 35: Descriptive Statistics for Uranium in Pond 7and Pond 10 (Pond 7 n=105 and Pond 10 n=134)...... 328 Table 36: Redox reactions in pond bottom sediments (adapted from Reddy et al., 1986 and in Avnimelech and Ritvo, 2003)...... 334

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Chapter 1: Introduction

1 INTRODUCTION

1.1 Introduction

Over fishing of the world’s oceans and a global increase in the demand for seafood has led to a worldwide shortage. To keep up with the demand of the food source, countries such as Australia have looked towards other means of supplying seafood for consumption both here and overseas (Burford et al., 2003). Aquaculture provides a source of marine and freshwater organisms for human consumption and reduces the impact of over fishing.

Most of the world’s marine aquaculture is undertaken in coastal regions in the tropics and sub-tropics. These areas are farmed due to their optimal climatic conditions and the proximity to seawater. These areas are generally estuarine, have established mangrove communities, and sustain fragile and complex ecosystems. In Australia there are strict environmental restrictions on the clearing of mangroves in estuarine ecosystems (Preston et al., 2002). However, uncontrolled prawn farming, in regions such as Asia, Central America and South America has led to the clearing mangroves for the excavation of ponds (Naylor et al., 1998).

There are three main types of prawn cultured in Australia. They are: Penaeus monodon (black tiger prawn), Fenneropenaeus merguiensis (banana prawn) and Penaeus japonicus (kuruma prawn). These three species support a $70 million dollar industry with production of over 4,000 tons annually (www.apfa.com.au). The average yield in Australia is about 4,500 kg/ha. Globally, the main prawn producers are Asia, the Americas and the Middle East. The world prawn production is in excess of 4 million tons annually (http://www.foodmarketexchange).

Estuarine environments are commonly associated with Acid Sulfate Soil (ASS) (Sammut et al., 1995; Sammut et al., 1996). Acid sulfate soils contain naturally formed pyrite (FeS2). Under waterlogged or anoxic conditions, the pyrite is non-reactive. When earthen ponds containing ASS are excavated for marine aquaculture, the pyrite oxidises and releases both sulfuric acid and metals into the receiving water. The release of sulfuric acid lowers the soil pH to less than

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Chapter 1: Introduction

4. At this pH and under newly oxic conditions, heavy metals that are associated with deposited minerals in the ASS dissolve and release metals into the surrounding soil and water (Dent, 1986; Sammut et al, 1996). Through bioaccumulation, farming in an ASS affected environment has the potential to impact the food chain. The ions that are released from ASS are harmful to farmed species and estuarine environments receiving discharge waters. Losses in production due to ASS impact are costly to Australia’s prawn industry and the economy.

1.2 The Research Problem

Previous work on ASS’s has been published; however there is a lack of knowledge on the hydrochemistry of the water in aquaculture ponds and the chemical relationship between ASS and aquaculture processes.

This project is designed to show the impact on pond water chemistry when ponds interface with ASS: farm practices that are designed to optimise production exacerbate the problem by driving oxidation reactions and accelerating reaction rates. Previous studies have considered the effects of ASS on the water chemistry in fishponds (Seo and Boyd, 2001). These studies are not suitable for comparison with prawn aquaculture for reasons that will become apparent in this thesis. Past studies have not adequately characterised the spatio-temporal chemical variation in ponds influenced by ASS.

Through the careful design of a permanent sampling network; the collection of regular samples throughout the growing season; the analysis and interpretation of these data, the author will provide unambiguous evidence that:

• pond water quality is influenced by the host soil

• the chemical impact is dependent on:

• the proximity of the soil

• the duration of exposure

• farming practices

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Chapter 1: Introduction

It is theorised that soil/water interactions create unsuitable conditions for the commercial production of sediment-dwelling prawn species such as the Kuruma Prawn (Penaeus japonicus: the species grown at Tomei during this study). By identifying and describing the association between ASS and commercial prawn farming, this study endeavours to suggest ways of reducing the effects of ASS on culture species and the surrounding environment.

1.3 The Study Setting and Location

This study investigates the relationship between water chemistry and acid sulfate soils (ASS) using spatial and temporal data from two semi intensive prawn aquaculture ponds. These ponds are located near Pimpama in the Woongoolba region in southeast Queensland and are owned by Tomei Pty Ltd (Figure 1).

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Chapter 1: Introduction

Figure 1: Location Map for the Woongoolba coastal flood plain and the Tomei Farm (marked in red)

1.4 Woongoolba/Pimpama Region - Geology and Literature

The Woongoolba region, as defined by Lohe (1980), is located on the Beenleigh Block of the New England Orogen. The Beenleigh Block consists of folded and faulted Late Palaeozoic to Early Mesozoic metasediments of the

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Chapter 1: Introduction

Neranleigh-Fernvale Group. These rocks form the “basement” which underlie (and provided some of the detritus to form) unconsolidated Late Pleistocene to Holocene sediments. This sedimentary veneer contains authigenic pyrite (Preda and Cox 2000). The estuarine geomorphic setting that accommodated the deposition of this sedimentary veneer is associated with the formation of ASS (Dent, 1986). The geological setting is described more fully in Chapter 2.

Other notable studies relating to the Pimpama and the Woongoolba region were undertaken by Preda and Cox (1998b); Preda (1999); Preda and Cox (2000); Cook et al., (2000); and, Preda and Cox (2001).

1.4.1 Research on Pimpama and Woongoolba region

Preda and Cox (2000) confirmed that the authigenic sedimentary pyrite associated with the ASS at Pimpama was deposited during a mid-Holocene transgression. Anthropogenic activities have exposed these sediments to the atmosphere which has caused the sulfide minerals to oxidise and hydrolyse. The oxidation of pyrite produces Fe and sulfuric acid according to Equation 1:

2+ - Equation 1 FeS2  Fe + S

The direction of this reaction is dependent on:

1. Redox conditions; the rate is dependent on environmental variables like rainfall, temperature, humidity and salinity.

2. Land use and land management practices play a crucial role in determining both redox state and reaction rate.

The oxidation of pyrite enables the breakdown of metal-ion bearing sediments and can lead to intense leaching of major and minor trace elements into local surface waters (Preda and Cox, 2001).The sulfate in the sedimentary veneer is derived from the Pimpama catchment seawater (Preda, 1999).

Preda and Cox (2000) suggest that the inorganic oxygen reactions can be initiated or sped up by micro-organisms such as Thiobacillus. This generally occurs at a pH of 4 where these bacteria increase the rate of oxidation of ferrous to ferric iron and hence the rate of acid production.

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Chapter 1: Introduction

During a dry season in Pimpama, the water table drops and sediment is exposed to oxygen. This is where minerals such as pyrite (FeS2), are oxidised to form goethite (FeOOH), jarosite (KFe3 (SO4)2(OH)6), gibbsite (Al(OH)3) and hematite (Fe2O3). Once a storm event or water infiltrates the sediment and reaches the oxidised zone, the water drains from the soils, discharging Fe3+, 2- + SO4 , H , Al and other trace metals and elements.

1.4.2 Research at Tomei

Govinnage (2001) undertook a study on the effect of ASS on prawns. She found that the Fe2+ released during pyrite oxidation was oxidised to Fe3+ on the gills of the prawns. This in itself is not a problem until it bonds with the oxygen being produced by the prawn’s osmoregulation. Once this occurs, Fe oxide forms and blocks the sites on the prawn’s gills for oxygen absorption, and the prawn suffocates.

Gosavi (2004) studied the soil and water chemistry associated with pond dyke walls at Tomei. That author confirmed the presence of pyrite in the pond dyke walls and concluded that its distribution was heterogeneous. Gosavi (2004) also recognised active pyrite oxidation in the dyke walls as the trigger for soil acidification (pH was measured to be as low as 3-4.5). This was attributed to rainfall seepage into the dyke structure. This low Total Dissolved Solids (TDS), oxygenated water led to pyrite oxidation and the release of Fe, Al and Mn into the adjacent pond. That author concluded that the pond dyke walls, and the inner core, are a significant source of acidity to the pond. Gosavi (2004) noted that most of the acidity related to ASS resulted from secondary reactions such 2+ 3+ as the oxidation of Fe to Fe , and the formation of ferrihydrite (Fe(OH)3): not pyrite oxidation alone as erroneously reported by other authors.

1.5 Aims

The study described in this thesis was undertaken to achieve the following aims:

• To confirm whether the ponds studied were chemically homogeneous or heterogeneous

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Chapter 1: Introduction

• If the ponds are heterogeneous, then to identify the spatial, temporal, physical and chemical characteristics in the water column (including pore water in sediments)

• To identify if the ASS contribute to the chemistry of the ponds

• To understand and build a framework to describe the relationship between chemistry in the pond water and soil

• To capture and collate data describing the pond environment

• To propose modifications to management practices to minimise the impact (if any) of ASS on the farms productivity

• Suggest ways of remediation and reclamation of ASS in similar environments

1.6 Hypotheses

The following hypotheses were tested to satisfy the aims of the study:

• The ponds are not chemically homogenous at any stage of the growth season

• Over the duration of the grow-out season, the water quality of the ponds would deteriorate

• Stratification of the water column would be more pronounced with depth and time due to the proximity to the chemical source and ongoing reactivity of the pond walls and base

• The water samples would contain concentrations of trace constituents that are found in the sediment (ASS)

• Sediment dwelling species (especially P. japonicus which burrow) are not the best type of species to use for aquaculture in this particular environment

• Any exposure of ASS to the atmosphere will lead to further leaching and exacerbate this environmental problem.

1.7 The Database

The database upon which the study is based comprises 238 water samples as summarised in Table 1. Twenty two sediment core samples were taken from

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Chapter 1: Introduction the two prawn ponds at Tomei Pty Ltd, Pimpama, Queensland.          

                 ! Data collected from the two ponds are presented in the thesis; however the ponds are not for direct chemical comparison due to their differing mineralogy.

There are 3 samples (and replicates) collected from three sites outside the farm (Figure 2) and are to be used as the chemical background or reference samples. The author collected all the water and soil samples. The samples are referred to throughout this text using the identification system 7/1-3: Pond 7, piesometer 1, outtake 3; or likewise; 10/1-3: Pond 10, piesometer 1, outtake 3. A summary of the accompanying data available to the author is summarised in Table 2.

Table 1: Water sample database Sample Type Location Pond Number Total Collection Date Pond Water Water column 7 24 Nov-01 27 Feb-02 24 75 Apr-02 10 24 Nov-01 25 Feb-02 23 Apr-02 21 93 Jun-02 Sand pore water 7 5 Nov-01 5 Feb-02 5 15 Apr-02 10 5 Nov-01 5 Feb-02 5 Apr-02 5 20 Jun-02 ASS pore water 7 5 Nov-01 5 Feb-02 5 15 Apr-02 10 5 Nov-01 5 Feb-02 5 Apr-02 5 20 Jun-02 238 Seawater 20m from farm intake/discharge N/A 1 Feb-04 Reference point in channel 50m from junction of farm channel N/A 1 Feb-04 and estuary junction between estuary and N/A 1 Feb-04 ocean at Jumpinpin Bar

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Chapter 1: Introduction

Figure 2: Sample location map for the reference seawater samples taken from the Moreton Bay - Jumpinpin Bar Estuary System Table 2: Summary of data used in this project

Category Description Source Cultural Published maps 1:100,000 Topographic Map 1 Air Photographs 1944, 1970, 1990, 1995 and 2002 2 Digital data Meteorological data (from three weather station 3 locations) Location map from Google Earth 4 Geological Unpublished geological map Map of Neranleigh-Fernvale Group 5 Published geological maps Numerous maps from published papers (see refs) 6 Sedimentological Soil Analysis – LECO/LOI LECO (Sulfur/Nitrogen/Carbon) and Loss On Ignition 7 (Organic Carbon) Soil Analysis XRD – dominant minerals as identified by interpretation 8 of the sample scan Hydrological Field measurements Water levels referenced to the permanently installed 9 multilevel piesometer Chemical Physical Variables Temperature, Electrical Conductivity, pH, Eh, 10 Dissolved O2 and CO2 + + 2+ 2+ 2+ 3+ - Major and trace ion chemical Major ions: Na , K , Ca , Mg , Fe and Fe : HCO3 11 - 2- 2- data ,Cl , SO4 and S Trace ions: Ag, Al, As, B, Ba, Be, Bi, Cd, Ce, Co, Cr, Cs, Cu, Dy, Er, Eu, Ga, Ge, Ho, La, Li, Lu, Mn, Mo, Nb, Nd, Ni, P, Pb, Pr, Rb, Sb, Sc, Sm, Sn, Sr, Ta, Tb, Th,

U, V, W, Y, Yb, Zn, Zr. 1 Department of Land and Water Conservation - Map Centre

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Chapter 1: Introduction

2 Natural Resources and Mines - Queensland 3 Bureau of Meteorology, Brisbane 4 Google Earth (www.earth.google.com) 5 Lohe (1980) Unpublished PhD thesis – see reference list 6 Numerous Sources – see reference list 7 UNSW School of BEES – XRF Laboratory – samples analysed by Mrs Irene Wainwright 8 UNSW School of BEES – XRD Laboratory – samples analysed and interpretation by Dr Irvin Slansky 9 Measurements recorded by the author 10 Samples collected and analysed by the author 11 Samples collected by the author and analysed in the UNSW School of Geography and Geology chemical laboratory 1.8 Project Outline

The primary focus of this thesis is to study the pond water chemistry and determine its chemical composition and define its chemical evolution. The author has approached the study holistically and integrated the water data with sediment analysis. This approach best depicts the interaction of the components of this system. The structure of the thesis takes on a similar format. Chapter 1 introduces the study and highlights its importance; Chapter 2 introduces the study environment; Chapter 3 gives a background to ASS; Chapter 4 presents an overview of previous work on water chemistry of environments affected by ASS; Chapter 5 is a literature review associated with aquaculture; Chapter 6 describes the management regime of the farm; Chapter 7 is a summary of the methods used to collect samples and techniques of analysis; Chapters 8 presents findings from this study related to the chemistry of the ponds sediments; Chapters 9, 10, and 11 are an in depth discussion related to the pond hydrogeochemistry; Chapter 12 discusses hydrogeochemical processes occurring in the water column and pore water sediment; Chapter 13 contains the summary of findings and presents a brief set of conclusions and recommendations; Chapter 14 contains the references; and Chapter 15 contains the appendices.

1.9 Summary

Prior to this study very little work has been undertaken on detailed hydrogeochemistry in aquaculture ponds either in Australia or overseas. This study has led to the collection, analysis, characterisation and interpretation of the chemical reactions that consistently take place in the pond water column,

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Chapter 1: Introduction in the pore space of the adjacent sediment and furthers the understanding of how these environments interact: as such, this study is a significant contribution to geochemical and aquaculture research, and the industry as a whole.

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

2 ENVIRONMENTAL SETTING – TOMEI FARM, SE QUEENSLAND

2.1 Introduction

This chapter provides information about the location of the study site, the climate, geological setting, geomorphology and anthropogenic activities that have and continue to influence the soil and water chemistry of the ponds at Tomei.

2.2 Location and Physiography

The Tomei farm is located in the Woongoolba region in south east Queensland. The Woongoolba region is a large coastal plain with most of the area being within 5m of sea level. The Tomei farm is adjacent to the Moreton Bay – Jumpinpin Bar estuary system and is between the Logan and Pimpama Rivers. The Tomei farm is located approximately 30km south of Brisbane (6932000mN, 534000mE) adjacent to cane farmland and an estuary system that connects Moreton Bay to the Pacific Ocean (Figure 3).

Figure 3: Woongoolba region showing Tomei Farm Location and adjacent waterways

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

Moreton Bay covers an area of approximately 1400km2. Its widest part is in the north covering about 35km and narrows in the south to about 5km (Lang et al., 1998).

The farm is located in an estuarine environment and covers an area of approximately 20 hectares (Plate 1). A total of 20 ponds are stocked with Penaeus monodon (P. monodon) and Penaeus japonicus (P. japonicus) post larvae (PL’s) at the beginning of each grow-out season. The ponds were excavated in a mixture of estuarine sediment (including ASS) and sand/clay material (ancient sand dunes located at the western sector of the farm). Dyke walls separate each pond and were constructed using a mixture of excavated pond sediments and capped with road base from a local quarry. A mangrove system separates the ponds from the estuary. The dominant mangrove species is Avicennia marina. The ponds are hydraulically interconnected to the estuary by an artificial channel. Tomei management uses the water from the estuary for pond water exchanges (intake and discharge).

Plate 1: Air photo of the Tomei Farm showing layout of the aquaculture ponds, intake and discharge points and adjacent seawater sample locations (Natural Resources and Mines Qld, 2002)

Two ponds were selected for the following reasons:

Pond 7 is adjacent to the mangroves and was flagged by farm management to have low annual productivity (based on results from previous years of farming).

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

The pond was selected for analysis as it was agreed upon by farm management and the researcher that it would be the typical ASS affected pond (due to the dominant acid sulfate clay type). Gosavi (2004) identified Pond 7 as a “high risk” ASS pond based on production data from 1998-2000 and soil acidity testing (including Total Actual Acidity (TAA) and Total Potential Acidity (TPA)) (Figure 4).

Pond 10 is distal to the mangroves, is underlain by sandy-clay sediment and has a history of better annual prawn production. It was proposed as a strong contrast to Pond 7. Pond 10 was identified by Gosavi (2004) as a “low risk” ASS pond (Figure 4).

As these two ponds were identified by previous work as a high and a low risk pond, this author postulated that they would provide good chemical comparisons between each of the ponds.

Figure 4: Historical data for Ponds 7 and 10, (a) Prawn Production 1998-2000, and (b) averaged Total Actual Acidity (TAA) and Total Potential Acidity (TPA) from 2001 (modified from Gosavi, 2004)

2.2.1 Evidence of ASS in the Ponds

Precipitated Fe oxyhydroxides and Fe monosulfide accumulation are important indicators of oxidised ASS. Red iron straining is common on the exposed pond walls and base at the end of a farming season. This colouration is due to precipitated Fe oxyhydroxides (hydroxides and hydrous oxides) both of which

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland are commonly associated with ASS. The pigmentation is most notable on the pond walls (Plate 2). The author visited the site at the conclusion of the 1999/2000 season and identified accumulated Fe monosulfides in the bottom of the ponds. Field Eh measurements of the sediment pore water taken at the end of the same grow-out season confirmed that the pond floor sediments were anoxic (Gosavi, 2004).

Plate 2: Iron oxyhydroxide staining on pond walls (a) and (b): this is a common feature associated with the oxidation of ASS when exposed to the atmosphere

2.3 Climate

The climate at Pimpama is subtropical with a wet summer and dry winter. Weather data for the three closest weather stations was provided by the Queensland Bureau of Meteorology and is presented in Figure 6, Figure 7 and Appendix 1.

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

Rainfall data from (1) Rocky Point Sugar Mill (40319); (2) Logan City Water Treatment (40854); and (3) The Gold Coast Seaway station (40764). The location of these stations is presented in Figure 5. Rocky Point Sugar Mill is the closest weather station to Tomei; unfortunately it only records rainfall data, so the other two stations were used to provide indicative climate data.

Figure 5: Location of the weather station – 1) Logan City Water Treatment Station, 2) Gold Coast Seaway and 3) Rocky Point Sugar Mill (modified from Queensland road map)

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

2.3.1 Rainfall

Summer is characterised by high rainfall, which often results in localised flooding. Preda and Cox (1998b) stated that about 70% of rain at Pimpama falls between November and April and it is commonly related to storm events.

Data collected at the Logan (station # 40854), Gold Coast Seaway (station # 40764) and Rocky Point Sugar Mill weather stations (www.bom.gov.au) are summarised in Figure 6.

Figure 6: Rainfall and temperature data summary for the three weather stations around the Tomei farm area (Gold Coast Seaway station #40764 and Logan WTS station #40854)

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

The rainfall data shows considerable variation between sites and months. The greatest variation in data is for the month of November. Rocky Point Sugar Mill is the closest to the Tomei farm and is deemed to be the most representative.

2.3.2 Temperature

The diurnal temperature data for the months during which field data were collected are presented in Figure 6.

The hottest month is typically February and the coldest is June. These two months also have the greatest range in air temperatures. The air temperature and solar radiation affects the temperature of the pond water. Fluctuation of water temperatures over a relatively short period of time, effects the pond water chemistry and in turn, the status of the prawn’s health. The shallow depth of the ponds (maximum is 1.5m) allows thermal stratification to occur. During the day when paddle wheels are functioning, this is not a problem, however when they are turned off in the late afternoon (generally about 5pm each day) the water does not mix and then stratified. The paddle wheels are not switched on again until about 7am.

2.3.3 Wind Patterns

Averaged 09:00hrs and 15:00hrs wind speed data for the Logan and Gold Coast weather stations show that wind speeds are greater in the afternoons (Figure 7). Records from the Gold Coast Seaway weather station (#40764) show that February and April have the highest average wind speed during the four sampling months, whereas at Logan this occurred in February. Wind speed is important to the ponds water chemistry because the greater the wind speed, the greater the potential for increased evaporation and diffusion of dissolved oxygen (DO) into the pond water column through wind sheer driven mixing. It is likely that on sunny days, algae in the ponds reach the greatest level of photosynthesis just after noon (when the ponds would be receiving the greatest intensity of solar radiation). Oxygen is produced during photosynthesis therefore on windy, sunny days the greatest amount of DO should be measured in the water column.

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

Figure 7: Morning and afternoon wind speed and solar radiation data summary (Gold Coast Seaway station #40764 and Logan WTS station #40854)

2.4 Geological Setting

The Woongoolba region is located on the Beenleigh Block of the New England Orogen: the most easterly tectonic element of continental Australia (Figure 8). The Beenleigh Block consists of Palaeozoic regionally metamorphosed sediment and volcanic units of the Neranleigh-Fernvale Beds (Cranfield et al., 1976). The green schist grade meta-sediments consist of greywacke,

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland feldspathic lithic and oolitic arenite, chert, jasper, shale, conglomerate, and intermediate and basic volcanic units. The Neranleigh-Fernvale Beds have an approximate minimum thickness of 7000m.

In southeast Queensland, the depositional drift direction is to the north. This is due to trade winds blowing predominantly from the southeast and from the wave swell from the ocean (Lockhart et al., 1998). In areas where shearing has occurred, tectonic breccias have formed. However, in other areas of the unit, the original structures such as bedding and relict pillow lavas/lapilli are still evident.

The rock sequence is most simply divided into two distinct groups:

• Basement - comprised of rocks from the Late Palaeozoic to Early Mesozoic Neranleigh – Fernvale Beds (Willmott et al., 1976; Preda and Cox, 2000) and

• Unconsolidated sediments – deposited during the Late Pleistocene to Holocene periods (Preda and Cox, 2000).

Generalised lithological comments for the Basement lithologies are presented in Appendix 2.

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

Figure 8: Tectonic Elements of the New England Orogen in Queensland: the study area is located on the most easterly tectonic subdivision - the Beenleigh Block (modified from Lohe, 1980)

2.4.1 Deposition of the unconsolidated sediments

Figure 9 presents the relative sea level fluctuation for the Australian craton over the last 150,000 years.

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

Figure 9: Relative sea level curve for Australia interpreted from the Papua New Guinea shoreline record: 150,000 years to present day (modified from Graham and Larsen, 1999)

The period from 10,000-3,000 yBP is when modern-day ASS formed after the last global transgression (Cook et al, 2000). This transgression caused river valleys and shallow basins to be flooded creating new estuaries and embayments. This oceanic influx created anaerobic conditions which through biologically mediated redox reactions resulted in iron bonding with sulfates (from seawater) and to form pyrite. This reaction was catalysed (biochemically) by the accumulated organic matter (Cook et al, 2000).

From 10,000-6,500 yBP, Moreton Bay was infilled in response to a global rise in sea level (transgression) (Neil, 1998). From 6,500-3,000 yBP, sea level in the Woongoolba region was approximately 1m higher than it is presently. Since 3,000 yBP, sea level has fallen to its current height (Preda, 1999 - from Flood, 1981).

2.4.1.1 The Formation of the Logan-Pimpama Coastal Area

Lockhart et al. (1998) stated that a marine transgression (advance of the ocean across former land) led to widespread flooding of the coastal plain, the drowning of river systems and the formation of present day estuaries. Lockhart et al. (1998) used data from continuous seismic profiling, drill hole data and maps of palaeo-fluvial systems to construct a coastal evolution model for the southern region of Moreton Bay (Figure 10). These authors concluded that the Logan-Pimpama coastal area underwent four sedimentation phases, they were:

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

Phase 1 – Late lowstand fluvial channel - when the sea level first started to rise and resulted in the deposition of 4-5m of gravel and sand. These sediments are generally confined to the thalweg of the valley and the valley walls.

Phase 2 – Early transgressive - the rising sea level created estuarine conditions and mud, silt and minor estuarine sands were deposited under reducing conditions. This phase is deposited over the first phase, without any evidence of an erosional surface. The sediments are olive green to black in colour and are generally deposited with pebbles and shells. Shells taken from the were dated and were aged between 13,650 ±50 yBP (at the base, -42m) and 7,430 ±50 yBP (at the top of phase 2 sediments, -20m). The sediments at the bottom of the depositional sequence were deposited a lot slower than at the top of this sequence. Phase 3 – Late transgressive to early highstand-open estuarine/marine conditions continue and with progressive sediment infill there was a progressive change to open shallow lagoonal deposition. Predominantly marine quartzose sands were deposited but these are known to contain lithic fragments (chert and metasiltstone), heavy minerals (ilmenite, zircon, rutile, and hornblende) and feldspar (plagioclase and orthoclase) with some reworked shell fragments. Phase 4 – Highstand – mixed fluvial/marine conditions with associated sediments varying from sand to peat.

Historical accounts report that Jumpinpin tidal inlet was reportedly closed until May 1898 when it was opened by extremely strong south-easterly gales (Lockhart et al., 1998; and Neil, 1998). Prior to 1898, when the inlet was closed, all sediments were derived from the Logan, Pimpama and Coomera Rivers (Lockhart et al., 1998). Lockhart et al. (1998) reports high sedimentation rates have been achieved during this period.

The mangrove communities that dominate the Moreton Bay estuary are believed to have taken hold in response to a change in environmental conditions (Neil, 1998). With sediment infill during phases 2 and 3 the clear seawater that previously supported coral communities would have become

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland turbid, as oceanic circulation was restricted. The coral communities then died and were replaced by more suitable organisms such as mangroves and sea grass.

Figure 10: Schematic cross section of the southern Moreton Bay Coastal Plain showing generalised lithologies and depositional environments (Tomei region is located near Jacob’s well) after Lockhart et al., 1998.

2.4.1.2 Pimpama River and tributaries in the area

The Pimpama River is located about 80km south of Brisbane and flows from west to east (Figure 3). This river (and its tributaries) drains an area of about 100km2 in the Moreton region and has an overall length of about 30km (Preda and Cox, 1998a). The river flows from the Darlington Ranges (Late Palaeozoic, Neranleigh-Fernvale Bed) at an elevation of 300m and falls away into a widened valley, then across a coastal flood plain, where it intersects its main tributary (Hotham Creek), then passes finally into and swamps, before discharging into southern Moreton Bay. During dry months, the flow is mainly from groundwater seepage, but during summer, the river flows strongly after heavy rain and flooding. Changes to the rivers flow have been made since urbanisation of the area with the establishment of sugar cane farms, infrastructure projects and housing. The flow has been modified by the use of tidal gates built on the lower coastal plains. Infiltration has been reduced by road building and housing development and this has led to an increase in runoff. The tidal parts of the river have measured salinities in the range of 20,000 – 30,000 mg/L and a pH of about 5. This decreases to pH 3 when rainfall flushes leachate from river bank sediments (Preda and Cox, 1998a). Acid is introduced into the Pimpama River by two main processes (Preda and Cox, 1998b):

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

Earthworks, pumping and drainage which exposes ASS to the atmosphere rainfall mobilising leachate after extended dry periods and ASS oxidation

These two processes have been further exacerbated by land clearing for industry and development.

2.5 Geomorphology

As described above, the global Holocene transgression played an important role in shaping the Moreton Bay estuarine environment. At the beginning of the Holocene transgression, the climate was warm and the polar ice sheets started to melt. Over time, the melted water caused a rise in sea level and the flooding of ancient valleys. Once the valleys were full of water, the overspill flooded into adjoining landscapes and estuaries (Graham and Larsen 2000a and 2000b). Approximately 6,500 years after the sea level stabilised to where it is today (Thom and Roy 1985; Ahern et al. 1999).

The progressive transition of depositional environments with the Holocene transgression led to the wide distribution of ASS along the south east Queensland coastline. The soil type commonly found in the Pimpama area is classified as a sulfuric Hydrosol (this is because it has a horizon or layer of AASS) (Isbell, 1996; Powell and Ahern, 1999). Acid sulfate soil layers are commonly buried beneath other soils of alluvial, colluvial or aeolian origin. Using the Isbell scheme, such soils are usually classified as Black or Brown Dermosols, Vertosols or Sandy textured Tenosols (Powell and Ahern, 1999).

2.5.1 Coastal Environments

Coastal environments are shaped by long term processes such as sea level fluctuation, tectonic activity, marine and/or fluvial sediment deposition and short term events, such as storm events (including flooding) and tides (Preda, 1999). Therefore, deposited sediment will undergo reactions depending on type of minerals deposited, mechanical and chemical influences during and after deposition, and climatic characteristics for that particular location. They are important areas for mineralisation of organic matter. This is attributed to microbiological activity, and in the case of the nitrogen cycle, (primarily) from benthic bacteria located at the sediment surface or deeper in the sediment via

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland root zones (Herbert, 1999). In many cases, estuaries and near-shore environments contain isotopes that suggest that some sediment has originated in a terrigenous environment (Jorgensen, 1983).

Many coastal environments contain ASS. The present day extent of pyrite oxidation depends primarily on anthropogenic utilisation of the land. Generally, two types of sediments are characteristically found in coastal environments. They are oxic sediments and anoxic sediments. Oxic sediments contain oxidised mineral species and aerobic microbes, whereas, anoxic sediment contains reduced mineral species and anaerobic microbes. The latter, are more abundant in this type of environment, as there are large amounts of organic material and sulfur which are biochemically altered with the end product being anoxic sediment.

As well as estuarine environments occurring in coastal zones, areas which are protected from the mechanical energy of the sea are termed tidal flats. These sediments are made up of reworking of marine sediments and/or sediment delivered by a fluvial source. Whatever the form of deposition, the sediments are fine-grained.

2.6 Recent History: Anthropogenic effects

2.6.1 Land use: Development of Aquaculture Land

During the 1960s and 1970s, a tidal flood mitigation scheme was established to assist with the development of the agricultural industry (sugar cane and fruit plantations). Drains and tidal flap gates were constructed along the Pimpama and Logan Rivers to restrict tidal flooding of the coastal plain (Preda and Cox, 1998a; and Preda, 1999). Reports written at the time suggested that this had the potential to create acid conditions.

Preda and Cox (2000) later commented that the surface and shallow ground waters in the Pimpama catchment have been affected by the construction of canals and drains. These structures have changed the hydrology of the lower catchment by cutting off meanders and changing the tidal regime by linking the Pimpama and Logan Rivers.

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Chapter 2: Environmental Setting – Tomei Farm, SE Queensland

Where Hotham Creek intersects the Pimpama River there is a chert quarry. The land around the area has been cleared for cattle grazing and sugar cane production (Preda and Cox, 1998a). High concentrations of pyrite were noted here by Preda and Cox (1998a).

These pyritic sediments have since been oxidised by changed exposure to weather, tides and land use. Anthropogenic activities such as quarry excavation, drain construction and land clearing directly exposes unoxidised sulfidic minerals or can cause the water table to drop. Aside of anthropogenic activity, rainfall is a natural mechanism for the chemical alteration of ASS. The amount of rainfall correlates with the amount of pyrite that gets oxidised and therefore the amount of sulfuric acid that is washed into the rivers. The pH of rivers containing acidic freshwater runoff is balanced when tidally influenced seawater is mixed through chemical buffering and mixing of the low pH water (Preda and Cox, 1998a).

Based on a review of the available air photos, little change has occurred to the surrounding land since the construction of the Tomei farm ponds in the early 1990’s. The land to the west and south west was utilised for cane farming, and the land to the north was covered by mangroves: this is still the situation today.

2.7 Summary

The prawn aquaculture farm studied is located on the coast at Pimpama, in south east Queensland, Australia. The geomorphology of the farm is attributed to the farms proximity to the coast and recent anthropogenic activity. The geology and subsequent deposited ASS are due to the depositional environment created by flooding during the last global transgression 10,000 yBP. Sediment was deposited on the flood plain from: the mountain range to the west of the farm; and, from reworked marine sediments from the east- southeast. The sandy clay soils and the proximity to the coast make the farms location favourable for the establishment of marine aquaculture. However, with the excavation of the farm sediments, exposure of unoxidised ASS has occurred. It has also meant that the existing mangrove colonies have been reduced in size and acidic discharge water from the farm enters the Moreton Bay estuary system.

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Chapter 3: Review: Focus on Acid Sulfate Soils

3 REVIEW: FOCUS ON ACID SULFATE SOILS

3.1 Introduction

Most of the literature related to ASS has focussed on issues in agriculture and urban environments. This chapter provides a background on the chemistry of acid sulfate soils particularly in relation to impacts on the natural environment and in aquaculture ponds.

3.2 Definition of Acid Sulfate Soils (ASS)

Acid sulfate soils are typically classified as having a pH<4, are a result of iron disulfide oxidation, (predominantly in the form of pyrite: cubic FeS2) and develop in areas where the production of acid exceeds the neutralising capacity of the soil (Dent, 1986; van Breemen, 1988). Acid sulfate soil contains heavy metals (such as Al, Fe and Mn) and traces (As, Cd, Cu, Zn) (Preda and Cox, 2001, 2002).

Any reference to ASS in this thesis should be read to mean Actual Acid Sulfate Soil (AASS) as apposed to Potential Acid Sulfate Soil (PASS). Tulau (1999) characterise AASS as containing highly acidic soil horizons or layers. These acidic layers result from oxidation of sulfide-rich horizons often mainly consisting of pyrite (FeS2). Oxidation results from exposure to the atmosphere by excavation, disturbance or drainage and produces acidity in excess of the sediment’s neutralising capacity (Equation 2).

3+ 2+ 2+ + Equation 2 FeS2(s) + 14Fe (aq) + 8H2O(aq) bbb 15Fe (aq) + 2SO4 (aq) + 16H (aq)

This oxidation reaction typically results in soils having a pH of 4 or less. In contrast to AASS, PASS’s have not been exposed to oxygen, and sulfide minerals remain largely unoxidised. The in-situ pH for PASS is greater than 4 and may be neutral or slightly alkaline (Bowman, 1993; Tulau, 1999). The process of PASS becoming ASS simply requires a change in environmental conditions resulting in exposure to the oxygen.

During precipitation, metal ions (particularly Fe) bond to oxyhydroxides, hydroxides, oxides, sulfates and sulfides. Iron sulfide precipitation removes sulfur, iron and trace metals from seawater (Morse and Cornwell, 1987).

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Chapter 3: Review: Focus on Acid Sulfate Soils

3.2.1 Depositional Environments

Acid sulfate soils can occur inland as non-tidal scalds (associated with saline, acidic sulfate seeps); mine stockpiles or tailings commonly referred to as AMD or Acid Mine Drainage; in close proximity to sulfide-rich sediments; or more commonly in coastal environments such as tidal mangrove swamps.

Inland non-tidal scalds result from rising saline groundwater. Such groundwater forms a seep when it reaches the ground surface. As the groundwater passes through the underlying rock and soil it can mobilise sulfur, iron, clay and heavy metals and deliver these to groundwater discharge areas. In anoxic portions of groundwater discharge zones sulfide minerals will precipitate even when the pH-Eh conditions become favourable. If conditions subsequently reverse and these zones are oxidised, sulfuric acid is released and pH falls creating an ASS (Fitzpatrick et al., 1998). The accumulation of elements at the surface leads to chemical reactions that break down clay minerals, making them dispersive, and resulting in soil and vegetation scalding and erosion.

Precious and base mineral deposits are often associated with metal-rich host rocks (Fitzpatrick et al., 1998). When ore is excavated from a mine, and is exposed to the atmosphere sulfides oxidise, releasing sulfur, acid and iron. The resulting runoff is termed Acid Mine Drainage (AMD). Oxidation of the pyrite leads to the formation of minerals such as ferrihydrite, lepidocrocite, goethite, jarosite and schwertmannite (Fitzpatrick et al., 1998). In many cases, the runoff will not only contaminate land in the vicinity of the mine, but may be transported as a leachate plume several kilometres away from the point source, resulting in significant environmental degradation if not controlled.

Acid sulfate soils associated with coastal embayments contain sulfides, particularly pyrite. This pyrite can be sedimentary (deposited with the marine sediment) or authigenic (accumulated by reduction of soluble sulfates during or after deposition) (Brinkman and Pons, 1973). The ASS forms in environments that are either: below sea level; or by low sedimentation rates in tropical, humid environments (Pons and van Breemen, 1982; and van Breemen, 1988). In the tropics, ASS’s form in tidal mangrove swamps, in marshy back swamps or

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Chapter 3: Review: Focus on Acid Sulfate Soils lagoons, or in marine terrace deposits. Potential acid sulfate soils are deposited on coastal plains, and are subsequently overlain by alluvial deposits (Sammut et al., 1996).

In sedimentary environments, the common iron minerals that form are hematite, goethite, siderite, glauconite (marine) and pyrite (and the precursor monosulfides species; mackinawite and greigite). Fitzpatrick et al. (1993) summarised the common iron minerals associated with ASS, this is shown in Table 3.

Table 3: Oxyhydroxides, hydroxides and oxides, sulfides and sulfates from Fitzpatrick et al. (1993).

Mineral Formula Goethite a-FeOOH Lepidocrocite g-FeOOH Akaganeite b-FeOOH

Hematite a-Fe2O3

Ferrihydrite Fe5HO8.4H2O Feroxyhyte d-FeOOH

Pyrite, Marcasite FeS2

Greigite Fe3S4

Mackinawite Fe9S8

Jarosite KFe3(OH)6(SO4)2

Schwertmannite Fe8O8(OH)6SO4

Pyrite, Marcasite FeS2

3.2.1.1 Pyrite Precipitation

Pyrite precipitation is an anaerobic process: it is confined to the reducing zone in the sediments (Rickard, 1995). Authigenic sedimentary pyrite is formed in environments that contain sources of sulfate (often seawater), iron and organic material. Bacteria require organic matter as an energy source to enable the reduction of sulfate and to form pyrite. When excess Fe and S are present, the amount and type of organic matter present in the sediment determines how much pyrite is formed (Berner, 1970).

Berner (1984) noted that it is not the concentration of iron that is important in pyrite precipitation, but rather the form of that iron. Fine-grained hydrous ferric oxides, which normally occur as rust-coloured coatings on other mineral grains are conducive to the precipitation of pyrite. Other less reactive sources of iron

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Chapter 3: Review: Focus on Acid Sulfate Soils are contained within clay minerals and are present in unweathered primary minerals such as biotite, pyroxenes, amphiboles, magnetite, and ilmenite. These must first be released into solution by weathering so are not available for pyrite precipitation.

The principal source of sulfur for pyrite precipitation in marine sediments is from sulfate dissolved in seawater (Berner, 1970). Authigenic sedimentary pyrite precipitation occurs in anaerobic, waterlogged conditions such as those found in tidal flats, saline or brackish swamps. Acid sulfate soils containing pyrite are generally geomorphically located in low lying coastal areas and are associated with deposited Holocene marine clays and sands. These areas also contain mangroves and other euryhaline vegetation communities which provide organic material to drive bacterially mediated sulfate reduction (Equation 3):

2- - Equation 3 2CH2O + SO4  2HCO3 + H2S

Once H2S is formed, it reacts with various iron-rich minerals to precipitate as an iron sulfide (Berner, 1970).

According to Berner (1970), the precipitation of pyrite occurs in the following order:

• Organic matter and iron is deposited in low energy marine environments.

• The available oxygen is metabolised by sulfur reducing bacteria to form

hydrogen sulfide (H2S).

• Dissolved H2S reacts with iron to form monosulfides. If H2S formation ceases and no more elemental sulfur (S) is added, greigite, mackinawite, or pyrrhotite precipitate and little or no pyrite forms.

• If the sulfate reduction continues, the H2S concentration increases and

detrital iron is scavenged to form FeS. Because excess H2S is produced, it diffuses out of the sediment and is released into the 2- overlying water and oxidises to form SO4 .

• The H2S that remains in the sediment forms elemental sulfur.

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• The excess elemental sulfur reacts slowly with FeS in the sediment to form pyrite which crystallises into framboidal spheres (Meyer, 1977)).

Pons and van Breemen (1982) and van Breemen (1988) suggest that there are six “essential ingredients for the precipitation of pyrite, they are:

• a continuous source of sulfate (typically seawater but can be sulfate- rich groundwaters)

• sediments that contain reactive iron oxides

• metabolisable organic matter (CH2O)

• the presence of sulfate-reducing bacteria

• an anaerobic environment

• limited aeration for the oxidation of all sulfide to disulfide. van Breemen (1988) shows that pyrite precipitation requires sulfur reducing bacteria to reduce sulfate to sulfide; the partial oxidation of the sulfide to disulfide and then the resulting disulfide; to react with Fe oxides. In freshwater sediments sulfate availability usually limits pyrite precipitation (Berner, 1984).

Pyrite precipitation is an on going process that occurs whenever environmental conditions are suitable, such as present day mangroves.

The sedimentation rate influences pyrite precipitation. Berner (1984) suggested that low sedimentation rates result in minimal pyrite precipitation. Rapid sedimentation, in nutrient-rich waters, results in rapid burial and preservation of organic matter and reduces H2S out-gassing, resulting in more pyrite precipitation.

Organic matter deposited on the bottom of a water-body acts like an oxygen- consuming barrier, however dissolved oxygen in the water column is still able to diffuse into the sediment layer by molecular diffusion, bioturbation, or wave and current stirring (Berner, 1984). A few centimetres below the sediment- water interface, oxygen is no longer present (due to the consumption by oxic bacteria) and the sediment is anaerobic. The depth of the anaerobic zone depends on the local environment. Plant roots and burrowing organisms may push the oxic zone deeper into the sediment in that direct area and during this

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Chapter 3: Review: Focus on Acid Sulfate Soils mechanical process pyrite precipitation can only occur deeper in the sediment profile.

3.2.1.2 Pyrite Morphology

Pyrite crystals are generally found in ASS as cubes, small spherical framboids (Plate 3) or in larger crystals as elongated clusters. Pyrite framboids are often associated with organic matter and therefore are observed under a microscope in plant cells of dead plant material (van Breemen, 1988). However, during deposition and compaction of the sediment (by tidal action and wave mechanics), the framboids may be disseminated (Fitzpatrick et al., 1993).

The depositional environment of pyrite framboid formation can be just as important as the mechanisms of deposition when it comes to controlling the concentration of pyrite framboids in the sediment. Fitzpatrick et al. (1993) found that the tidal soils contained the highest concentrations of sulfidic materials, inland soils contained intermediate concentrations and mine soils studied, contained the least. An experiment by Morse and Wang (1997) suggested why there is variation between different environmental settings. Their experiment observed the growth of pyrite over time and under different pH conditions. They found that the formation of pyrite was influenced by pH and sulfide concentrations. The pyrite that formed under neutral to alkaline conditions had a small-sized (~2 μm) euhedron structure. Pyrite that was formed under acidic conditions took on a spherulitic shape and was dominated by various surface structures that resembled authigenic sedimentary pyrite framboids. The size and distribution of pyrite framboids are now being used as a tool to determine the depositional environment (Liaghati et al. 2005).

Goldhaber and Kaplan (1975) suggested that a single crystal of pyrite may form if there is direct precipitation of the pyrite as opposed to the formation of framboidal pyrite. Single crystals of pyrite and framboids are generally found in clay and silt-sized sediment fractions, or as infillings of foraminifera and diatoms. Goldhaber and Kaplan (1975) suggested that framboids are produced by the replacement of aggregates of organic cells with pyrite. When the framboid spheres are in groups, they are termed polyframboids.

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The mineral marcasite, is stoicheiometrically identical to pyrite, but has only been found in some ancient sediment, it has not been identified in recent marine sediments (Goldhaber and Kaplan, 1975).

Plate 3: Scanning Electron Microscope (SEM) image of framboidal pyrite which is the common form associated with organic matter (from Preda and Cox, 2004); (a) fresh pyrite crystals, (b) Framboid affected by acidity over several months of exposure. Many crystal edges are etched and covered with iron phases.

3.2.1.3 Pyrite Oxidation

Oxidation of pyrite occurs when pyritic sediments are exposed to a source of oxygen and water (Equation 4 – note that this is the reverse of Equation 3).

- 2- Equation 4 2HCO3 + H2S 2CH2O + SO4

This may occur during the drainage of land; disturbance of soil due to infrastructure works, urban and tourism development, or mining. Anthropogenic oxidation of ASS has led to the most extensive environmental damage. This process can occur naturally (Golez and Kyuma, 1997). In the mangrove environments, organisms such as the mound building mud lobster (Thalassina anomala) enhance natural oxidation of ASS during their excavation process. Similarly, prawns such as Penaeus japonicus, which bury themselves in the sediment, have an effect on oxidising reduced sediments. The pyrite in ASS can also be oxidised by the roots of plants. As they push deeper into the soil profile, they form macropores which introduce small amounts of oxygen (Blunden and Indraratna, 2001). Burrowing macroinvertebrates also facilitate the oxidation of sediments (Figure 11). The activity of these organisms can cause anoxic environments to be exposed to

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Chapter 3: Review: Focus on Acid Sulfate Soils oxygen and create ASS. It may also cause a flux of nutrients and heavy metals from ASS into the water column (Palmer et al., 2000).

Figure 11: Worm burrows in ASS facilitate the oxidation of sediments (modified from Kristensen, 2000)

Pyrite oxidation results in a surface coating of ferric sulfate on sediment grains in acidic conditions; when alkaline conditions prevail the surface coating consists of iron (lll) oxyhydroxide (goethite) (Todd et al., 2003). The rate of oxidation is pH dependent: below a pH of about 5, Fe2+ oxidation is relatively slow but once the pH decreases to about 3.5-4.0, the oxidation process is catalysed by Thiobacillus ferrooxidans (Bloomfield, 1973). These bacteria are active at low pH’s and assist in accelerating the rate of pyrite oxidation by 5-6 times and therefore, the rate of acid production (McKibben and Barnes, 1986; Ahern et al., 1999; Preda, 1999). Thiobacillus ferrooxidans optimally grows when the temperature is about 30°C and at a pH of about 3 (Ahern and McElnea, 1999; Cook et al., 2000).

Thiobacillus thiooxidans oxidises elemental sulfur, making the soil acidic, which promotes the presence of precipitated Fe3+ to catalyse the oxidation of pyrite. In the presence of these bacteria, the amount of Fe3+ is increased by a factor of 20 (Bloomfield, 1973).

The initial process of pyrite oxidation is represented by the following equation (Dent and Bowman, 1996; Hey, 1999; Cook et al., 2000):

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Equation 5 FeS2 +15/4 O2 + 7/2 H2O  Fe(OH)3 + 2H2(SO4) (Stage 1) pyrite + oxygen + water  ferric hydroxide + sulfuric acid bacteria

In the complete process, ferrous iron (Fe2+) is oxidised to ferric iron (Fe3+) and 2- 2- sulfide (S ) is oxidised to sulfate (SO4 ).

The final stages of pyrite oxidation are represented by the following equations:

The Fe-S bonds are broken and pyrite is oxidised to sulfuric acid:

2+ + 2- Equation 6 FeS2 (s) + 7/2 O2 (g, aq) + H2O  Fe (aq) + 2H (aq) + 2SO4 (aq) (Stage 2)

Ferrous iron is oxidised to ferric iron:

2+ + 3+ Equation 7 Fe (aq) + 1/4 O2 (g, aq) + H (aq)  Fe + ½ H2O (Stage 3) Thiobacillus ferrooxidans

At pH <4, Fe3+ can remain soluble and can react as the oxidant with the pyrite to produce further acid.

Ferric iron precipitates to form ferrihydrite:

3+ + Equation 8 Fe + 3 H2O  Fe(OH)3 + 3H (Stage 4)

During the fourth stage, if the pH is greater than 4, the precipitation of iron hydroxide releases more acid. This reaction is called secondary oxidation or hydrolysis.

The oxidation of sulfur phases to sulfate also occurs over a number of stages. This is due to the large number of electrons (14) that are transferred during the oxidation process. The microbes contribute to the process as they extract the sulfate from seawater and reduce it to sulfide (Graham and Larsen, 1999).

Preda (1999) suggested that pyrite oxidation consists of three the processes:

• an initiation phase triggered by oxygen;

• an acid-generating phase when sulfate starts to form and;

3+ • a catalytic phase when the hydrolysis of pyrite in the presence of Fe is catalysed by bacterial activity.

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Whereas Rimstidt and Vuughan (2003) describe the stages of pyrite oxidation as:

• the cathodic reaction;

• electron transport; and

• anodic reaction.

During the cathodic reaction, electrons are accepted from Fe(ll) on the mineral surface. The oxidation process oxidises sulfur and not the iron. The iron remains as Fe2+ as it is released into the solution. Electrons are transported from the anodic site to the cathodic site and then to the oxidant, this is an eight step process by which sulfide becomes sulfate. Sulfur atoms undergo several oxidation stages during this final stage of the anodic reaction and the transport of electrons is thought to happen one at a time. It is possibly the reason for the presence of so many sulfur compounds in the process (Rimstidt and Vuughan, 2003).

The oxidation of ferrous iron (the initial product of pyrite oxidation) to iron + hydroxide consumes O2 and releases hydronium ions (H3O ). This process decreases DO concentrations and the pH (Sammut et al., 1995). The ions H+, Fe (ll) and Al are the dominant cations involved in ASS runoff, where H+ = the 2+ acid source, Fe = from reduced O2 in water, Al = toxic (to aquatic organisms), although other heavy metals are also associated with ASS oxidation.

Cook et al. (2000) found that the acidity of drainage water from ASS was due to both proton (H+) and metal acidity. The metal acidity was generated when metals underwent hydrolysis and oxidation reactions. These reactions commonly occur when the ASS runoff mixes with receiving waters that have an alkaline pH.

Previous to Cook et al. (2000), the common method for identifying ASS was measurement of the pH alone. However, titratable acidity or analysis for metal ions such as ferrous iron and aluminium is a better indication of the acidity of drainage water from ASS.

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High iron concentrations in the water have the potential to reduce oxygen in runoff. The H+ is a form of acidity, however metal acidity attributed to the ions Fe2+ and Al3+ also contribute to the total acid budget of ASS’s (Cook et al., 2000).

Minor elements such as As (9.6 wt. %), Co (2.2 wt. %), Sb (0.7 wt. %), Au (0.3 wt. %) and Ni (0.2 wt. %) can be associated with the pyrite lattice and are released during pyrite oxidation. Trace elements such as Ag, Bi, Cd, Hg, Mo, Pb, Pd, Pt, Ru, Sb, Se, Te, Tl, and Zn are also thought to be present in the mineral lattice associated with pyrite (Rimstidt and Vuughan, 2003).

Secondary minerals can form during the oxidation of pyrite. This is generally related to pH. Iron becomes fully oxidised and the pH reaches 3.7, jarosite

(KFe3 (SO4)2(OH)6) can form (Preda, 1999). This is a yellow precipitate (Plate 4) and is formed according to the following reaction (Equation 9):

+ + Equation 9 Fe(OH)3 + H2SO4 + K  KFe3(SO4)2(OH)6 + 3H2O + H

Jarosite is the potassium end-member of the jarosite-alunite group (Baron and Palmer, 1996) and in the above equation; potassium is mainly contributed to the reaction by silicate sources such as clays (Fitzpatrick et al., 1993).

The oxidation of pyrite is detrimental to aquatic ecosystems particularly to flora and fauna that are not mobile. Leaching and discharge of ASS into rivers and estuaries impact on fish, crustaceans, annelid worms, shellfish, and oysters. The latter four organisms are comparatively less mobile than fish and are generally unable to escape the ASS runoff.

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Plate 4: SEM images of Jarosite (a) synthetic Jarosite, (b) naturally occurring Jarosite crystal taken from www.tltc.ttu.edu and www.earthsciences.uq.edu.au

3.2.2 Sulfides and ASS

Over the last million years, sulfur has been accumulating in the ocean in its reduced form, however it is mostly present as the oxidised form, sulfate. All plants, animals and bacteria in the ocean use dissolved or ingested sulfate for metabolising proteins (Goldhaber and Kaplan, 1975; Meyer, 1977). Sulfate is believed to have a residence time in the ocean of about 40 million years and pyrite has been found in sedimentary rocks that are over 250 million years old. Dissolved sulfate makes up about 10% of rocks such as limestone, shale, and metamorphic rocks; and the concentration is a little less in granite (8%), sandstone (7%), and basalt (5%). Sulfate has a direct influence on the calcium and magnesium cycle in both water and in sediments.

Morse (1995) suggested that the following reactions occur during sedimentary sulfide formation:

2- • organic matter + SO4 bacteria  H2S

• reactive Fe (such as mackinawite or greigite) + H2S  Acid volatile sulfur (AVS)

• H2S + oxidising bacteria  elemental sulfur (So)

• AVS + So  pyrite

The above reactions suggest that for formation of pyrite to occur, a source of sulfur is required. Sammut et al. (1996) stated that iron sulfides that are found

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Chapter 3: Review: Focus on Acid Sulfate Soils associated with ASS are generally in the form of pyrite (cubic FeS2). They also attribute seawater to being the source of sulfur. The seawater was in contact with the deposited sediments during the Holocene and the sulfate would have been reduced by microbes to form sulfide. These microbes generally belong to the Desulfovibrio sp. and use organic acids or H+ for the reduction of sulfate. Methane formation does not seem to occur in marine sediment until sulfate reduction is complete. This may be due to two factors: (1) sulfate-reducing bacteria compete favourably for H2 liberated during fermentation, or (2) H2S is toxic for methane bacteria. Sulfate reduction continues over periods of several million years and to depths of several hundred metres within the sediment column. The production of sulfide is closely related to the availability of organic matter.

Re-oxidation of reduced sulfides in estuarine sediment occurs rapidly in the upper sediment layers (van Breemen, 1992; Holmer et al., 1994). This is due to bioturbation (by burrowing crabs and worms), plant root interference or from tidal mixing introducing oxygen into the anoxic system. Generally the deep, anoxic sediment sulfur is locked-up in different forms of precipitated iron sulfides (Holmer et al., 1994). Rapid sedimentation of marine sediments excludes oxygen, causes bacterial sulfate reduction and leads to the production of high concentrations of H2S (Berner, 1984). Plants are very sensitive to H2S and even at low concentrations, it can impair root functioning and increase the plants susceptibility to infection (Dent, 1986).

3.2.2.1 Iron monosulfide Formation

Iron monosulfides are the intermediate step during the formation of pyrite. They are generally formed in waterlogged environments such as drains, shallow coastal lakes and swamps when there is a lack of aeration in the sediments on the bottom of the water body, and complete pyritisation of sulfide does not occur (Pons and van Breemen, 1982). Once they are exposed to the atmosphere, they rapidly oxidise. Two main monosulfides are amorphous FeS known as mackinawite ((Fe, Ni)9S8) and greigite (Fe3S4). Mackinawite is a metastable iron-sulfur phase and is an important precursor to pyrite precipitation in sedimentary and hydrothermal systems. It is the first-formed iron sulfide phase following the reaction of ferrous ions with sulfide ions.

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Greigite occurs as black accumulations of crystals; often associated within the root remnants, has a tabular morphology and ranges in size (Bush and Sullivan, 1997) (Plate 5). It will rapidly oxidise at room temperature in the presence of water and is formed when there is a source of iron, sulfur and organic material (the latter used as an energy source for bacteria).

Plate 5: Scanning Electron Micrograph (SEM) of colloidal greigite (magnetic iron sulfide) taken from www.earth.leeds.ac.uk

3.2.3 ASS – The Distribution of ASS

There are approximately 12 million ha of ASS in the world (Beek et al., 1980 in Bronswijk et al., 1993; Bronswijk et al., 1995; Blunden and Indraratna (2000) (Figure 12). There is some uncertainty of the quantity of ASS in Australia as not all of the affected areas have been identified.

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Figure 12: Global distribution of ASS (modified from www.cimmyt.org)

Powell and Ahern (1999) estimated that there are about 2.3 million ha of ASS along the coast of Queensland covering a distance of 6, 500km. However, White et al. (1997) and Blunden and Indraratna (2000) stated that in Australia there are about 3 million ha and are thought to consist of about 1 billion tonnes of pyrite.

Naylor et al. (1995) mapped ASS in New South Wales and estimated that there are 0.4-0.6 million ha of ASS.

When completely oxidised, each tonne of the pyrite produces about 1.6 tonnes of sulfuric acid (Fitzpatrick et al., 1998; Tulau, 1999).

Clearly there is disagreement of the data in the literature, but what is apparent is that this is a widespread issue: there is a significant area of ASS’s affected terrestrial and aquatic ecosystems across the globe.

Acid sulfate soils have a similar impact where they are located in the world, they:

• Damage waterways making aquatic organism (fish, oysters, prawns) more susceptible to disease and mortality. This has been attributed to the lowering of water pH and release of Al into the water which

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damages epithelial cells and allows pathogens and fungus to invade the organism (Plate 6) (Sammut, 1998).

• Damage agricultural land and crops by impeding root growth due to aluminium toxicity; in severe cases, toxicity from ferric iron or manganese, low cation-exchange capacity due to preferential acid dissolution of smectite and illite; extreme acidity; temporary water logging and drowning of terrestrial plants in poorly drained, clay soils; precipitation of phosphate due to soluble aluminium; low concentrations of macro-nutrients (e.g. Cu and Mn); and young seedling dieback from

H2S or ferrous iron toxicity (Brinkman and Pons, 1973; Clark et al., 1996) (Plate 7);

• Corrode infrastructure due to acid release from sulfide oxidation (Clark et al., 1996; Hey, 1999)( Plate 8);

• Encourage the swamp mosquito (which seeks out acid drainage for breeding) (Clark et al., 1996);

• cause detrimental health problems (including stunted growth, mental issues, dermatitis, arsenocosis) through the release of metals such as Al and As into drinking water (Buchet et al.,1994; Clark et al., 1996; Buchet et al.,1996; Suner et al., 1999).

Plate 6: Yellow arrow points at a fungus (Aphanomyces invadans) attacking epithelial fish cells (Sammut, 1998)

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Plate 7: Acid sulfate soil exposure resulting in the death of vegetation down slope (Hey, 1999)

Plate 8: Corrosion of concrete pipes (Hey, 1999)

3.2.3.1 Australian Examples and Continued Awareness

The initial recognition of ASS in Australia (in the Macleay River Floodplain) is attributed to Walker, 1963 (Graham and Larsen, 1999; Walker, 1963). In 1972 Walker published his completed review on the subject. The First National Conference on Acid Sulfate Soils was held in Coolangatta in 1993 (Powell and Ahern, 1999). After this conference the Queensland Department of Natural Resources and Mines (NR&M) and the Queensland Acid Sulfate Soils Investigation Team (QASSIT) produced a map which delineated the extent of ASS between Logan and Coomera (Gold Coast region). Figure 13 presents a subset of this map over the Tomei prawn farm and adjacent areas.

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Figure 13: Subset of the Queensland Government (Natural Resources and Mines) Acid Sulfate Soils Map for the Logan – Coomera region

On the east coast of Australia, most of the ASS work has been carried out in New South Wales (NSW) (Preda, 1999). ASS maps have been constructed for NSW by Naylor et al. (1995) and these show that there are 600,000 ha of ASS in NSW. In Queensland (QLD), the Acid Sulfate Soil Investigation Team (QASSIT) has led the way in the study of ASS (Figure 14).

In the 1990s, ASS were identified in dry inland areas (see Fitzpatrick et al., 1996) in the Mt Lofty Ranges, South Australia.

In 1998, the Australian Acid Sulfate Soil’s manual was published. It outlined a set of guidelines for the assessment, management, laboratory methods and drainage guidelines (Stone et al., 1998).

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Figure 14: Queensland ASS distribution map

The identification of ASS affected land is assisted by consideration of the following factors (Hey, 1999):

On Land:

• Elevation: ASS’s are generally associated with low-lying, coastal areas with an elevation below 5m Australian Height Datum (AHD).

• Vegetation: Common species that are indicative of ASS are mangroves (marine environment), saltwater couch (salt tolerant), tea trees – Melaleuca sp. (water logged soil), she oaks – Casuarina glauca, and Phragmites sp. (salt and acid tolerant) (Hey, 1999). However, the lack of vegetation on the soil surface is also an indication of ASS. Acid scalds may occur on the soil surface as the acid can damage plant roots by stunting their growth or by creating an environment that is too acidic for the plant to live, causing a die back and areas void of vegetation.

• Cracking of soil: ASS’s once drained and dried out, shrink and crack, forming a mosaic pattern. Sometimes the soils can shrink to about 50%

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of their volume (Fitzpatrick et al., 1998). Once the soils crack, they often do not go back to their original state even with the application of water. ASS’s in marine environments contain sodium, and this makes them highly dispersible. When the soils are re-flooded, they tend to erode and disperse. This is a problem in both natural and artificial environments. Infrastructure, such as roads and slabs under houses built on ASS can become unstable and crack, causing structural damage (Plate 9 and Plate 10).

• Sulfidic odours: The soil or pore water from ASS can emit a sulfuric smell that is described as being like “rotten egg gas”. The smell is from the hydrogen sulfide gas being expelled during the oxidation process. In the mining industry, wastewater industry and in pits dug in ASS, the odour can prove to be fatal if exposed to gas without ventilation.

• Jarosite (KFe3(SO4)2(OH)6: this is an indicative mineral of pyrite oxidation. It is generally seen as yellow mottled deposits in old root channels, on river banks, or in soil cracks that have been exposed to air (Dent, 1986)(Plate 11).

Plate 9: Marine sediments have a poor load bearing capacity (from Hey, 1999)

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Plate 10: Drying out of marine sediments leads to shrinking and cracking resulting in a mosaic, prismatic pattern on the surface and extending deep in the soil profile (from Hey, 1999)

Plate 11: An indicative mineral of oxidised ASS is jarosite (from Hey, 1999)

In Water:

• Fish kills: Callinan et al. (1993) and Sammut et al. (1995) stated that epizootic ulcerative syndrome (EUS) is linked to acid sulfate discharging into estuaries and rivers. This not only has an impact on native fish species, but also on commercial fish stocks. Many of the fish kills that have been reported in the last 5-10 years have been attributed to ASS’s (Hey, 1999) (Plate 12). This is due to the export of acid, Al and Fe from the soils directly into the water bodies. If fish can, they will try to avoid the acidic water by swimming away, however, sometimes flood gates will stop them from escaping the acidic water. This leads to disease or mortality. Immobile aquatic organisms are unable to escape the acidic conditions, leading to their mortality. It is important to note however, that fish kills are not always related to ASS discharge. Sometimes they occur because of nutrient runoff from farms, extreme

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changes in temperature, increases in suspended sediment and organic loads, and low DO from decomposing organic material in the water (Hey, 1999).

• Iron: Red-brown flocs (iron hydroxide or oxyhydroxide) have been observed in waters at a pH above 2.7. These suspended flocs can be transported downstream or can coat stream beds and banks. Above pH 6, white flocs, possibly aluminium hydroxide or hydroxysulfate can form when there is mixing between acidic fresh water and brackish water (Sammut et al., 1996). During the process of oxidation and hydrolysis, the formation of Fe and Al flocculants smother the benthic biota (Cook et al., 2000). Dissolved reduced Fe in the water can become oxidised once it reaches the osmoregulating gill slits in fish and prawns (Simpson et al., 1983; Govinnage, 2001). This causes iron oxides to clog gills and suffocate the animal. If Fe is present in the water in its oxidised form, it may attach itself to algae or suspended sediment and smother vegetation and the benthic component. This can alter the chemistry and physical properties of the water; and may occur at the point source of ASS runoff, or in the case of a river regime, may be carried downstream to be deposited there (Sammut et al. 1995; Powell and Ahern, 1999).

• Infestation of Lyngbya majuscula (or fire weed) blooms in Moreton Bay (Qld, Australia) have been linked with elevated concentrations of iron in the water. Lyngbya majuscula effects human health by causing dermatitis, asthma and eye irritation and if inhaled or ingested, respiratory and digestion tract irritation. It also has caused fish and sea grass mortality (Hey, 1999) (Plate 13).

• pH: acid water in estuarine and river environments reduce the amount of alkalinity in the water, which is crucial for the growth of crustaceans (Sammut et al., 1996). In freshwater systems associated with ASS, the pH can drop below 2. Hey (1999) suggested that in areas the pH may be 4.5-5, and the acidity is not always related to ASS, but can be related to the tannins and organic acids which occur naturally in that type of environment.

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• Colour and clarity: An indicator of aluminium in a water body is the clear blue-green colour and unusual clarity (Plate 2). This is generally seen at mine sites, but also associated with ASS. Acid from ASS oxidation breaks down the structure of aluminosilicates found in clay. This process releases Al from the soil into the water and flocculates clay particles in the water, depositing them onto the bottom of the water body. Sammut et al. (1999) suggest that at a pH 5.2, Al toxicity to fish is at its highest due to the chemical species present at this pH.

• Corrosion: can be seen in concrete structures and shells in ASS environments. Acid that is discharged from oxidised ASS corrodes the calcium carbonate in the concrete (Plate 14). Over time, the problem becomes so great that bridge pylons, drains and support beams need to be replaced. Steel supports inside the concrete may also corrode and affect the structural integrity. Replacing infrastructure is a costly exercise however if left undetected it could prove to be dangerous. Hey (1999) accounted the experience in the Tweed Shire (NSW), where the local council had spent $4 million on the replacement of steel and concrete pipes which were found to contain large cracks due to the acid attack. The Maroochy Shire Council (QLD) spent $1 million replacing 1km of storm water drains (the pipes were expected to last 80 years, but lasted only 20 years. In the Hastings and Manning Rivers the integrity of oyster shells were compromised by acidic water discharging into oyster leases and degrading oyster shells. The water dissolved so

much CaCO3 that perforations formed in the shell. The perforations compromised the shell structure (which is natural protection for the oyster) and allowed acidic water to enter the shell. The acidic water caused changes in valve movements, decreased growth and reduced feeding rates. Over time, the acidic water and associated Fe accumulation caused large oyster mortality. This in turn, had and still has huge implications to the oyster industry (Dove, 2003).

• Change in ecology of water bodies: ASS’s draining into rivers and streams can cause the natural vegetation and fauna to die (Sammut et al. 1995). Acid tolerant species such as reeds, water weeds and lilies, often take the place of natural flora. In some cases, the acid tolerant

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plants can grow so prolifically on the water surface, that benthic plants may not receive enough light and oxygen to be able to grow, causing a change in the flora (and in turn fauna) at the bottom of the water body. On the other hand, water that receives ASS discharge may become clear due to aluminium causing flocculation. In this case, the plants and animals in the benthic zone, may receive too much light. This can cause stress or in some cases plants will grow more prolifically, they will diffuse so much oxygen, that the water will become saturated. This effect can cause fish kills as the fish receive too much oxygen (Sammut, 1996).

Plate 12: Major fish kill in Pimpama River in 1996 (from Hey, 1999)

An important precursor for the oxidation of pyritic material in ASS environments is the draining of land for agriculture and urban development (Powell and Ahern, 1999). Draining of ASS’s have been and are still for the construction of canal estates, housing/industrial estates, marinas, roads, golf courses, aquaculture ponds, sand/gravel extraction, and draining of land for sugar cane farms. Rassam and Cook (2002) went further stating that drainage; rainfall; irrigation and evapotranspiration (of crops such as cane which is a large evapotranspirator) are the main mechanisms for the liberation of acid and heavy metals from ASS. Cook et al. (1999) studied the soils in two areas of cane farm land in Pimpama, Queensland. One of the paddocks they obtained samples from had been drained for 30 years and the dominant sulfide mineral was jarosite. The drainage water associated with this site contained high concentrations of Fe and lower concentrations of Al. The other paddock sampled had been drained for just 5 years and predominantly contained pyrite.

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This paddock’s drainage water contained higher concentrations of Al than Fe. This study highlights the long term problem of metal and acid leaching once the soils have been exposed.

Plate 13: (a) Lyngbya majuscula floating in Moreton Bay, (b) Proliferation of the species shown at low tide – this is associated with Fe leaching out of adjacent ASS

Plate 14: Corrosion of concrete pylon in the Pimpama River (from Hey, 1999)

3.2.3.2 Overseas Examples

Acid Sulfate Soils are generally found in countries that are situated in subtropical or tropical environments such as Thailand, Malaysia, Vietnam and Central Africa. However, it is also a problem in countries such as England, The Netherlands, Scandinavian countries and New Zealand (Preda, 1999).

Bachmann et al. (2001) reviewed a situation in Germany, where acid soils are a problem, however this acidification is associated with an acidified mine lake.

In Finland ASS’s originated from artificially drained Holocene marine and lacustrine sulfide-bearing sediments (Astrom, 2001).

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Two English studies attribute acidification of rivers to sulfidic sediments (peat and ore material respectively). Soulsby (1995) noted that there was a relationship between large events of rainfall and subsequent acidification of stream water in Wales, UK. Acid and aluminium runoff occurred after a storm event and caused the mortality of macroinvertebrates and salmon. The Humber estuary, in England, consists of sediments deposited in the Holocene as in the case of Australian ASS’s. There, the river acidification is also attributed to exposed sediments from the closely situated mined areas and the clearing of land (Rees et al., 1998).

3.2.4 ASS and its impact on the aquatic environment

Seawater has a neutralising capacity of about 2 moles m-3. Therefore seawater buffers the pH (hydrogen ion) of acidic water, but will not decrease the metal acidity (as shown in the results in Chapters 10 & 11).

The exposure of ASS in freshwater environments (as in the case of agriculture or by rainfall), enables pH’s to drop, in some cases to negative pH values as there is not the buffering capacity of seawater (Sposito, 1998). Discharging of sediments from aquaculture production into rivers, estuaries and the ocean causes pollution and is not environmentally sustainable (Avnimelech and Ritvo, 2003).

During the dry season the process of oxidation is more intense and the production of acid is more obvious (as the water can contain around 15 mg/L of Al and 10 mg/L of Fe). Rainfall after a long period without rain washes the accumulated Al and Fe into rivers as an acidified leachate. This ASS leachate lowers the receiving waters pH and creates heavy metal toxicity for fish and other aquatic organisms (Preda and Cox, 1998a).

3.2.4.1 Estuaries, mangroves and associated land development

Acid sulfate soils are associated with mangrove environments and estuaries (Lin and Melville, 1992; Lin and Melville, 1993; Callinan et al. (1993). Vegetation such as mangroves, assist in the moderation of the water table through evapotranspiration and are important in the estuarine ecosystem. Mangroves are also important as they assist in stabilising banks and

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Chapter 3: Review: Focus on Acid Sulfate Soils decreasing the rate of erosion, they take up nutrients, produce large amounts of organic material that can be exported out to sea, and they provide habitats for juvenile crustaceans and fish (Loneragan et al., 1997). Mangroves also are able to accumulate heavy metals from ASS environments.

Preda and Cox (2002) undertook a study on trace metals in sediments and mangroves in the Pumicestone region (about 30km north of Brisbane). They identified trace metals in the sediments and pneumatophore tissue from mangroves (Avicennia marina) growing in the estuary. They found that their samples contained Cu, V, Cr, Zn, Fe and Mn. There were however differences in concentration, Cu was elevated in the mangrove tissue. Elements like Fe, Mn and Zn appeared to be present at fixed peak concentrations. Preda and Cox (2002) attributed this to the chemical reactions occurring at the sediment- water interface, the response of the plant to salinity and metal speciation.

A study by Lin et al. (1995) concentrated on an undrained floodplain, largely unimpacted by human activity. They found that due to raising and lowering water tables (in high rainfall and drought times); pyrite oxidation occurred naturally and influenced the non-pyritic topsoil layers by capillary mechanisms (Lin et al.,1995; Lin et al.,1998). Exposure of the pyrite leads to oxidation and production of acidic runoff containing iron and sulfate (Lin and Melville, 1992; and Callinan et al., 1993, Lin et al., 1996). Mangroves have been extensively cleared from estuarine environments both in Australia and around the world to free up land for development and farming. The clearing of vegetation such as mangroves disturbs large volumes of ASS’s and provides a mechanism for the exposure of buried anoxic pyrite.

Large rain events do not always produce a discharge of acid, particularly following long dry periods when storage capacity is high; however they can cause an initial slug of low pH runoff containing high concentrations of dissolved Al and low concentrations of oxygen to estuarine waters (Easton, 1989; Sammut et al., 1995; Hyne and Wilson, 1997; Melville et al., 1999; Willett et al.,1993; Wilson et al.,1999).

Acid sulfate soil drainage water (especially after major rain events) has been associated with an ulcerative fish disease, epizootic ulcerative syndrome

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(EUS) causing large fish mortality (Callinan et al., 1993). Sammut et al., (1995) studied the impact of acidification from ASS on an estuarine environment. Powell and Ahern (1999) found that fish start to show signs of stress at a pH<5.5. Sammut (1998) found that damage to gills occurs below pH 5.5 due to aluminium, whereas at pH 3 it is the acid that causes damage because aluminium at this pH is not toxic. Sammut et al. (1999) stated that humic acid and silica in the water would decrease the toxicity of aluminium under weakly acidic conditions.

It’s not only fish that are affected by ASS runoff. Sammut et al. (1999) discussed the effects ASS has on oysters. They found that oysters closed their valves when the pH dropped to 4. The valves were reopened when the pH was increased to above 7.5. Other impacts on oysters due to acidic water are reduced growth rates, shell bleaching and brittleness, poor quality of the oyster meat and oyster mortality (Dove, 2003).

3.2.4.2 Effects of ASS on rivers

In coastal regions of Australia, particularly in northern New South Wales and Queensland, acidity in rivers is attributed to acid sulfate soils. Land clearing is an important contributor to the acidification of rivers. Floodplains flanking rivers are usually high in nutrients and make for good farming land. Land reclamation and draining exposes PASS and oxidises it.

An article in a fishing magazine written by Easton (1989) described an event that occurred in the Tweed River, northern New South Wales. After water and aquatic tissue sampling, it was determined that the fish mortality was caused by acidic water discharging into the river from surrounding ASS. The sudden flushing of acidic discharge into river ecosystems from ASS’s and fish kill occurred in the Pimpama River, in November 1995 (Powell and Ahern, 1999). Infrequent storm events tend to mobilise most of the acid and heavy metals from ASS. Acid water does not discharge into rivers and creeks after every storm event particularly if the storm event is not the first for the season (Sammut et al., 1996; Russell and Helmke, 2002). Sammut et al. (1995) described that the discharge from ASS into rivers acts as a “chemical barrier to fish migration”. Sammut et al. (1996) study on the Richmond River (NSW,

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Australia), found that in river water which had a pH of 2.7, red-brown flocculants were formed and had been transported down stream or had coated the river bed or banks. Iron oxide/hydroxide staining of the soil on the banks of the river is a useful tool in ASS area identification.

3.2.5 ASS and Aquaculture

A number of studies have been carried out to determine the effects of pH on aquaculture environments. Singh (1981) compared the effects of the changing chemistry of ASS and neutral soil in brackish water fishponds in the Philippines. The study found that upon flooding of the ponds, the neutral soil took 3-4 weeks for the pH to stabilise, where it took 9 weeks for the pH to stabilise in the ASS affected pond. He suggested that this was an indication that the reduction processes were much slower in the ASS ponds.

Singh (1981) also concluded that the quality of the pond water is strongly affected by the nature of the pond bottom sediment particularly when the pond is located in ASS’s.

Pond dyke walls contribute a large percentage of acid to a pond during a rainfall event following dry conditions (Brinkman and Singh, 1982). Acid generated from pond sediments were correlated in several studies with chronically low yields to extensive mortality of culture organisms (Simpson and Pedini, 1985; Russell and Helmke, 2002).

Corfield (2000) studied the effects of acid sulfate runoff (ASR) on the subtidal estuarine macrobenthos in northern NSW and referred to the work undertaken by Alongi (19998) that correlated ASS and the reduced aquaculture productivity in Southeast Asia.

A detailed discussion of the interaction of ASS’s and aquaculture conditions is continued in Chapter 5.

3.2.6 Trace elements

3.2.6.1 Trace elements in Sediment pore water

Trace elements are generally contained in the weathered material of deposited rocks. They are considered to be present in the following forms; water soluble,

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Chapter 3: Review: Focus on Acid Sulfate Soils exchangeable, carbonate bound, ferric and manganic oxide bound, organic matter and sulfide bound, silicate bound, and residual. Absorption and desorption of trace elements generally occurs on the surface of clays, organic matter, and iron oxides in the sediment and are dependent on changes in the ionic composition or Eh-pH.

Trace elements are also bound to various insoluble organic forms such as living organisms, detritus, and humic material and can be present in the lattice structure of minerals (Beizile et al., 1989). As weathering reactions are very slow in clays, the trace elements tend to be locked up; therefore the trace element cycling is generally contributed to oxides and oxyhydroxides of iron, manganese and sulfides of iron (Luther, 1995). Iron and manganese oxides are particularly good at scavenging trace elements that are affected by sediment Eh and pH change. Iron and manganese can exist as nodules and concretions, cemented between particles or on particle coatings in sediments (Guo et al., 1997), and can act as temporary carrier phases for trace metals and can precipitate when there is a source of oxygen (Ouddane et al., 2001).

Sediments serve as sinks or sources of toxic heavy metals to water bodies. The exchange of metals across the sediment-water interface depends on the physical, chemical, and biological processes that control speciation, solubility, and the development of concentration gradients. The elements contained in pore water sediments never reach chemical equilibrium (Pardue and Patrick, 1995).

Petersen et al. (1995) simulated in the laboratory, the exchange of trace metals at the sediment-water interface. They used cores extracted from the tidal region of the River Elbe, Germany. They found that heavy metals such as Cd and Cu were dominant in the top 3mm of the core and suggested that this was because the area was associated with the particulate organic substances and therefore the Cd was able to be remobilised. Heavy metals such as Co and As were found to be soluble in the post-oxic and sulfidic zones and were associated with the oxide minerals. Cobalt was more enriched in the region where manganese reduction occurred and As was associated with the highest concentrations of Fe(II). Petersen et al. (1995) concluded that Co was

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Chapter 3: Review: Focus on Acid Sulfate Soils associated with Mn oxides and As was associated with Fe(II) oxides. They also suggested that As could be transported from iron reduction regions via diffusion.

Simpson et al. (1998) investigated the re-suspension of trace metals from anoxic sediments. They found that, particularly in estuarine environments, metals such as; Cu, Cd, Fe, Mn, Ni, Pb and Zn, were re-suspended into the overlying water during oxidation of sulfides. Once these metals are released from their sulfur species, they may be co-precipitated with Fe and Mn hydroxides or bound with organic matter.

Liaghati et al. (2003) studied the heavy metal distribution in sediments in the Bells Creek catchment, in southeast Queensland. They emphasised the importance of studying trace elements in soils as they are attributed to the contamination of surface water, shallow groundwater and bioaccumulation of the elements in the food chain. Heterogeneous trace metal accumulation was attributed to non-uniform weathering of the sediment.

Under reducing environments, such as those associated with buried sediments, trace elements are released into sediment pore water. The sediment pore water can be discharged into the local environment, transporting the trace elements from their source. Bowman et al. (2000) studied the East Trinity site, in Cairns and found that the concentrations of dissolved Al, Fe, and Zn in drainage waters often exceeded ANZECC Guidelines for aquatic ecosystems. The site was a former estuary and was drained for sugar cane production and therefore contained sulfidic mud’s (ASS). The site exported acid, heavy metals and As into the adjoining estuary.

Dissolved metal sulfides have been found to be the major sources of mobilised metal fractions. Astrom (1998) investigated the mobility of transition metals (Ti, Zn, Ni, Co, Mn, Fe, Cu, V and Cr) in ASS undergoing different stages of oxidation in Finland. He found the Mn, Zn, Ni, Co and Cu are released and mobilised in large concentrations upon oxidation of ASS.

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Metals that are not generally associated with pyritisation such as Pb, Zn and Cd, have faster water exchange reaction kinetics than Fe2+ and form MeS phases prior to FeS formation and subsequent pyrite precipitation. Metals such as Co and Ni are incorporated into pyrite during formation as they tend to have slower reaction kinetics than Fe2+ (Morse and Luther, 1999). van Breemen (1992) recognised that trace metals such as Ni and Co, may be present in the pyrite structure by either substituting for Fe or as related sulfides such as Cu, Zn, Pb and As. Therefore, oxidation of pyrite will also release associated trace metals into the environment.

Once trace metals are dissolved in the water, they can form precipitates such as metal carbonates or metal sulfides. The latter are formed in anoxic sediments, where bacteria reduce sulfate to sulfide. The sulfides can then react with iron and dissolved trace metals (such as Cd, Zn, Pb, Cu, and Ni) to form metal sulfide precipitates (Batley, 2003). Sulfide precipitation is very important in mangrove sediments. Low redox potential in mangrove sediments and the presence of H2S leads to the precipitation of sulfide minerals.

Most of the trace element solubilities are affected by pH and Eh. The solubility of Zn increases as pH decreases (Drever, 1997). Copper is mobilised from iron oxides during acid dissolution (Golez and Kyuma, 1997). Calmano et al., (1993) studied heavy metals in contaminated sediments, particularly the way they are bound and their mobilisation under different pH and Eh conditions. They noted that heavy metals could be remobilised upon oxidation of anoxic sediments, using water as a transportation medium. They linked pH and Eh by stating that most metals are mobilised once the pH dropped to below 4.5, however redox conditions control the mobilisation of particular metals e.g. Cd, Hg, Pb, Zn, Fe, Mn and Cu. The adsorption of Fe(ll), however, is related to pH; adsorption increases with increases in pH. As the pH increases, the surface charge becomes more negative, and the adsorption capacity increases. When the pH is between 6 and 7, the iron oxide coating on the sediment grains is positively charged, and hence the adsorption of Fe2+ takes place against electrostatic repulsion. As the pH increases the hydrated oxide surface becomes less positively charged. When this occurs, there are more sites

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Chapter 3: Review: Focus on Acid Sulfate Soils available for Fe2+ adsorption and therefore its adsorption capacity increases (Sharma et al., 1999).

Just as the pH and Eh are important variables in determining the concentration of trace elements in the sediment pore water, so too is the concentration of dissolved oxygen. Calmano et al. (1990) commented on previous work on the composition of anaerobic and aerobic pore water. They highlighted the effects of oxygen on elements in Table 4:

Table 4: Composition of anaerobic and aerobic pore water (after Maab et al., 1985 – in Calmano et al., 1990 pg 509)

Ion Anaerobic pore water Aerobic pore water Nitrate <3 mg/L 120 mg/L Ammonium 125 mg/L <3 mg/L Iron 79 mg/L <3 mg/L Zinc <30 μg/L >5000 μg/L Cadmium <0.1 μg/L 80 μg/L Arsenic 150 μg/L 15 μg/L

Herbert (1999) stated maximum emissions from estuarine environments occur at night. That author related this observation to benthic microalgae, at night they do not produce oxygen, denitrification occurs and this enables the nitrous oxide to be released into the water column.

3.2.6.2 Arsenic and Acid Sulfate Soils

Arsenic is a metalloid whose mobility and availability is largely controlled in anoxic environments by sorption on sulfide minerals. Its toxicity is dependant on the dominant species. In the coastal plain environment, As can be derived from the following minerals: realgar (As4S4); orpiment (As2S3); arsenolite

(As2O3) and iron minerals such as arsenopyrite (FeAsS) and loellingite

(FeAs2). Arsenopyrite is the most common As bearing mineral.

Dissolved arsenic in natural waters is generally composed of the inorganic forms arsenate [As(V)] and arsenite [As(lll)] with the organoarsenicals dimethylarsinic acid (DMAA) and monomethylarsonic acid (MMAA) found in much smaller proportions. The amount of As released to the aquatic environment is related to adsorption reactions on mineral surfaces. Since the rate of As(V) adsorption is faster, it follows that more As(V) will be taken up by

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Chapter 3: Review: Focus on Acid Sulfate Soils the Fe (oxyhydr)oxide surface, resulting in the observed fall in the relative proportion of As(V) in solution. Manganese oxides and Fe (oxyhydr)oxides are known to be effective oxidants of As(lll) (Gault et al., 2003).

Trace elements are also absorbed by marine flora. The relationship between arsenate (As(V)) and phosphorus in sea grass leaves was tested by Fourqurean and Cai (2001). They found that the ratios of each ion were related to the amounts in the water where the sea grasses are growing. As(V) is the inorganic form of arsenic and the most toxic as it inhibits the absorption of phosphate in plants and animals. In phytoplankton, As(V) absorption is restrained by high concentrations of phosphate. Marine algae uptake As(V) and in the process reduce it to the non-toxic form, arsenite (As(III)). Fourqurean and Cai (2001) found that different species of sea grasses are able to absorb different amounts of As, depending if the environment contains sources of P or As, or not.

Lengke and Tempel (2003) discuss some of the problems associated with the mobilisation of As in countries such as Bangladesh, India, Taiwan, Argentina, United States, Canada, China, Mexico and Australia.

Once As reaches the sediment-water interface, it is typically removed from the pore water by adsorption onto iron oxides. Under strong anaerobic conditions, the dissolution of iron and manganese oxides leads to the release of trace elements (As, Co and Cr). Once oxygen enters the system, the precipitation of Fe occurs and this tends to remove trace elements.

3.2.6.3 Trace elements in Biota

There is a correlation between trace element concentrations in sediment pore water and organism mortality (Nordstrom, 1982; Sammut et al., 1995; Bufflap and Allen, 1995b). The effect of Al on fish in aquatic ecosystems was researched and published by Sammut and Callinan, 2000; Govinnage, 2001; and Dove, 2003. Coagulated colloidal hydrated Al has been observed to accumulate on the gill surface, causing the fish to produce mucous in response. The build up of mucous on the gill surface in turn suffocates the fish. This, together with acidic waters, and Fe accumulation on gill slits of fish,

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Chapter 3: Review: Focus on Acid Sulfate Soils prawns and oysters are the main reasons for morbidity in ASS effected environments

Humans are at risk of absorbing trace metals from the marine environment. Suner et al. (1999) undertook a study on total and inorganic As concentrations in fauna in a Spanish estuary. They found that the organisms living in the estuary sediments, contained higher levels of inorganic arsenic than those living in the water column. This was attributed to ingestion of particles in sediments during feeding. Human consumption of these organisms was main source of As in the human diet. Inorganic arsenic (As(lll) and As(V)) is the most toxic form of As to humans, once ingested. Heinrich-Ramm et al. (2002) note that for most of the world’s population, the main source of As is through the digestion of seafood. They state that in general, As concentrations in seafood are one order of magnitude higher than that found in drinking water or in air. In their study, arsenic species in sample seafood were identified and compared to the amount and the species type in human urine after digestion. They noted that the dominant As species in marine fish is arsenobetaine, not the toxic form of As to humans (carcinogenic to humans are tri- and pentavalent inorganic arsenic species).

3.2.7 The cost of Acid sulfate soils to the community

Acid sulfate soils degrade the environment by eradicating the more sensitive aquatic species and this reduces species diversity. This occurs through repeated acidic discharge into estuaries and rivers.

Not only are ASS’s costly to the environment, their effects on constructed structures also cost Australia millions of dollars each year (Hey, 1999). The acidic runoff associated with ASS can damage (or destroy) infrastructure (such as concrete or metal supports); delay coastal developments; damage or kill export species such as fish, oysters or prawns; and damage crops by acidification of their roots (Powell and Ahern, 1999).

Land subsidence is a common feature of ASS’s. They are generally composed of water saturated gel and mud. Their poor load-bearing capacity causes buildings or roads to subside. This phenomenon was highlighted during the 1999-2002 upgrade of the Pacific Highway by Abigroup Pty Ltd. During the

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Chapter 3: Review: Focus on Acid Sulfate Soils construction of the Yelgun-Chinderah deviation in the Tweed, ASS subsidence cost the NSW Government $350 million and was the most expensive single road project undertaken in regional Australia. The upgrade was only 29km in length (www.rta.com.au).

3.2.8 Remediation of Acid Sulfate Soils

Sutherland et al. (1999) suggest methods for treating ASS: neutralisation with lime; burial in an anaerobic environment, sluicing of fines; water table manipulation; capping; and oxidation of the soils and subsequent treatment of discharge waters. However the most common form of remediation of ASS is by the application of agricultural lime (CaCO3) to the oxidised soil. In some cases dolomite (CaCO3.MgCO3) or magnesite (MgCO3) are also used to neutralise acidic water (White et al., 1997).

Lime is either directly applied on the soil or mixed in to increase the in-situ pH. There are problems associated with this process. Lime is relatively expensive. Dent (1986) calculated a ratio of 3:1 (by mass) of calcium carbonate would be needed to neutralise pyrite sulfur. At Tomei, during the 1999-2000 season, granulated lime was applied to the bottom of the ponds. It was not mixed into the soil and after the spreading process; wind picked up the lime and covered the surrounding trees. Also, by adding too much lime to naturally acidic water, there is a risk of growth of toxic aquatic algal species such as cyanobacteria (blue-green algae) (Desmier, pers. com. 2003). In some situations, the use of aggregated lime may be ineffective over time. This is because Fe and Al oxyhydroxides adsorb to the lime aggregates in a thin coating that is often referred to as “armouring” (McConchie and Clark, 2000).

Another method of delivering lime to neutralise ASS is by burying lime aggregate in the soil. Cook et al. (2002) have trialled the use of lime slots at a cane farm effected by ASS at Pimpama, south east Queensland. Slots were dug into the soil to a depth of 0.9m and 0.2m wide parallel to drains. The lime was mixed with the soil at a rate of 40kg/m length of slot. The water quality in the drains was measured over time and the results showed that pH and the Fe concentration reduced over time, but Al concentration increased in the drainage water. Appelo and Postma (1999) contributed this effect to the

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Chapter 3: Review: Focus on Acid Sulfate Soils displacement of Al ions from cation-exchange sites by the Ca ions (from the lime). However, an increased concentration of Al in the discharge water is undesirable because high concentrations of Al are toxic to aquatic organisms.

McConchie and Clark (2000) suggested an alternative way of treating ASS runoff, by the use of seawater-neutralised bauxite refinery residue, later named Bauxsol™ (Lin et al., 2002). Bauxsol™ is applied by mixing it into the soil and in places where ASS can discharge into drains or rivers; it can be mixed with soil and used like a chemical barrier. McConchie and Clark (2000) stated that the modified bauxite residue has a soil reaction pH of about 8.6 and an acid neutralising capacity of 3.65 moles/kg. They claim that unlike lime, the bauxite mud can not be leached from the soil by rain or groundwater and is more cost effective. McConchie and Clark (2000) state that the mud would remove trace metals from any water in contact with it. They also stated that treatment by the bauxite residue was beneficial because it enhances the nutrient retention capacity of the soil and promotes plant growth, and it has little effect on the concentration of bicarbonate ions that may accelerate pyrite decomposition. They stated that the selective extraction data show that only a small proportion (<2% for most elements tested – As, Ba, Cd, Ce, Co, Cr, Cu, Ga, La, Nb, Nd, Ni, Pb, Rb, Sc, Sr, Th, U, V, Y, and Zn) of the trace elements in the red mud could be removed by compulsive exchange reagents and hence, it was very unlikely that future leaching could lead to environmental contamination.

Lin et al. (2002) stated that during their experiment on Bauxsol™, the results showed that when more Bauxsol™ was applied; there was a larger retention of metals in the soil: it seemed that Bauxsol™ retained metals such as Al, Zn and Cu in the soil. It seems that the product was beneficial as long as there was Bauxsol™ in the soil, however, once the product was flushed from the soil matrix and acidic conditions returned, the question is would there be greater concentrations of these ions released into the runoff? If Bauxsol™ had not been used, these metals would have been leached, however, they may have been leached from the soil over a longer period of time and therefore in a less concentrated form than with the use of Bauxsol™. McConchie and Clark (2000) stated that a higher proportion of the trace metals could be released by the treatment when using 10% nitric acid, but given the high neutralising

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Chapter 3: Review: Focus on Acid Sulfate Soils capacity of the red mud, there was little chance that acid leaching would increase environmental metal loads. This concept offered by McConchie and Clark (2000) is contestable as water with a low pH (such as that found in drains associated with cane farming as in the case at Tweed Heads) (Lin et al., 1998) has been found to also contain a high concentrations of trace elements and the study at Tomei has proven that even seawater-buffered ASS runoff can mobilise trace elements. Ward et al. (2002) also suggested using seawater-neutralised bauxite refinery residue (SNBRR) in conjunction with calcite to remediate ASS. They found that the application of these materials reduced the rate of sulfide oxidation and sustained a pH close to pH 7. This author suggests that more consideration of the effects of BauxsolTM application should be studied before use in the field.

Another common way of remediating ASS’s is by flooding areas of acid sulfate affected soil with seawater. As seawater has an alkaline pH (generally 8.4) it is used to buffer the acidic pH of acid sulfate soils. One problem with this type of remediation is salinisation of the surrounding soil. This has been observed in land adjacent to aquaculture ponds in Asia (White et al., 1997). McConchie and Clark (2000) stated that treating ASS with seawater to neutralise acidity, uses the buffering activity of Ca and Mg in seawater, however in their study, they failed to comment on release of heavy metals from ASS upon exposure to oxygen. It is true that seawater will buffer the hydrogen ion acidity and increase the pH of discharging water; however, it does not assist in reducing metal acidity which is an important consideration when treating ASS runoff.

Whalen et al. (2000) propose another method of neutralising ASS, which is particularly common in Asia, is with the addition of cattle manure. Bicarbonates and organic acids in the manure raise the pH of the soils to near neutral. However a possible side affect of using the manure would be that it would increase the nutrient loadings to the soil and discharge water and could prove detrimental to the receiving aquatic environment.

3.3 Summary

Potential acid sulfate soils are an international problem, especially when changes in environmental conditions have oxidised these sediments to

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AASS’s. This has led to the oxidation of sulfide minerals and the release of acid and heavy metals. Exposure of unoxidised pyrite is generally due to the excavation of land by anthropogenic activity. Once pyrite oxidation occurs, the soil leaches acid, sulfate, iron and associated heavy metals. During rainfall events, the acid and metals are mobilised and often are discharged to rivers, streams or estuarine environments. Aquatic organisms and manmade (concrete and metal) structures are negatively impacted. There still is not an agreed optimal method for the remediation of ASS affected land: ASS impacts are costly to both the environment and to the economy.

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Chapter 4: Review: Water Chemistry In ASS Effected Ponds

4 REVIEW: WATER CHEMISTRY IN ASS EFFECTED PONDS

4.1 Introduction

Exposure of ASS creates problems at both local and regional scales. At the ASS source, such as in an excavated pond, water that is exposed to the soil acts as a medium and transports by-products from the ASS into the adjacent natural and modified environments. Aquaculture farms are generally located adjacent to an estuary and vegetation (such as mangroves) is cleared for the excavation of ponds. This location is selected because water from the estuary can be used to fill ponds in the initial stages and to “freshen” the ponds (removal of waste products and the addition of oxygenated water) during the grow-out period. Water from the estuary is also used to replace evaporated water in the ponds. During pond water exchange, turbid water is discharged into the estuary. It is during this process where nutrients, acid, sulfur, iron and heavy metals; liberated from the ASS in the ponds, enter estuaries. Three major elements have been discussed in the literature review chapters (Chapter 3, 4 and 5); the soil, water and aquaculture environment. These are the three most likely impacted elements affected by ASS oxidation (applicable to this study) and provide a well rounded background to support the validity of this thesis.

4.2 Water

4.3 Sampling water (in lake/ponds)

There seem to be as many ways of sampling surface water as published studies, however, when designing any type of scientific study, it is important to use a standard method which best suits the experiment. Grab sampling seems to be a common method of taking a water sample in from a body of water, however, if a temporal study is to be undertaken, then this method would make it almost impossible to sample the same spot each time (Figure 15). Sayles et al. (1973) stated that up until their study on marine sedimentary pore waters, all previous pore water samples had been generally obtained by squeezing pore water out of the soil cores in the laboratory. An in-situ method of obtaining pore water from sediments was first visited in their study. They used an in-situ

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Chapter 4: Review: Water Chemistry In ASS Effected Ponds probe that extracted filtered samples of pore water from a profile reaching 1.5 meters into the sediment.

Figure 15: Grab sampling methods undertaken in rivers and lakes (from Bartram and Balance, 1996)

In Höpner (1981) three main methods for sampling interstitial water from fine- grained sediments were delineated: pore water extraction from sediments by squeezing or centrifugation (which was also the preferred method for Adams (1994), Bufflap and Allen (1995a)), extraction using a pore water sampling device or, a diffusion sampler. Höpner (1981) stated that pore water extraction is the most frequently used method but as it is done ex-situ as there is the possibility of the extracted water to change its chemistry due to oxidation of the sample. The second method Höpner (1981) mentions is to use a sediment pore water suction device and filter the pore water through tubes. This type of pore water sampling was first used by Sayles et al. (1973) for detailed micro- scale studies. Höpner (1981) states that this type of sampling allows a water sample to be obtained under the same conditions, as it is in-situ at a fixed site. The third method of sampling uses a diffusion sampler. A diffusion sampler is inserted into the sediment and left for the pore water to reach equilibrium with the undisturbed pore water. This process works on a dialysis principle and was first used by Hasslein (1976). The sampler is often termed as a "peeper" (Bufflap and Allen 1995b; Angelidis, 1997) (Figure 16).

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Figure 16: Peeper (dialysis) in-situ water sampler

Disadvantages of using the dialysis type of sampling for this PhD study are:

• Divers are needed to install this type of diffusion sampler and therefore would be suitable only in shallow water environments (Hasslein, 1976; Adams, 1994)

• For anoxic sampling, the sampler should be prepared in an oxygen-

depleted environment so as not to introduce O2 to anoxic sediment. The preparation of the dialysis sampler takes a substantial amount of time and

ensuring the peeper remained O2 depleted during transportation and instillation was impossible

• The actual sample drawn from the dialysis chamber is of about 4 ml (Hasslein, 1976), so the sample size limits the type and amount of analyses the can be done on that sample.

• Anoxic pore water needs to be extracted from the chamber without being exposed to an oxic environment i.e. in a nitrogen glove box. This would mean that extraction would have to be done in an environment with a different temperature than the sample site. This means that there is a high probability of thermodynamic alteration of the sample and therefore an untrue representation of the pore water chemistry.

• The equilibration time of the peeper can range from one week to one month and in some studies, researches may not have this time or

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economic resources to wait (Hasslein, 1976; Bufflap and Allen, 1995a and 1995b).

• Erroneous data can occur from using the peeper. They may suffer from incomplete equilibration, membrane break down due to microbial activity, contamination or from the development of an electrical potential across the membrane, interfering with the dissociation of ions (Bufflap and Allen, 1995b; Angelidis, 1997)

• Due to the close proximal nature of the chambers in the diffusion sampler, the vertical profile obtained from the sampler would be on a very small scale.

Bufflap and Allen (1995b) suggest that the main problem with the in-situ samplers was they were not able to sample to the depth that some researches wanted. Their solution to this was the use of multilevel samplers. This was also the method preferred by the author for the water collection used in this thesis.

There are three important factors to consider when dealing with pore waters that are undergoing oxidation-reduction, adsorption-desorption and precipitation-dissolution processes (Angelidis, 1997). They are:

• the number of analyses that are carried out on the samples; as it will expose the sample to numerous atmospheres and temperatures, which is especially detrimental for the concentration of elements associated with the carbonate system;

• temperature; and

• time

If any of these three components are extended, then the concentration of the elements in the sample is likely to change and the sample would become unrepresentative (Angelidis, 1997).

Continuous sampling has become a popular method for the collection of water samples. Jarvie et al. (2001) undertook a study on a Scottish River, where they collected data and samples using a continuous water monitor. Conductivity, pH and temperature were measured every 15 minutes and spot samples were

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Chapter 4: Review: Water Chemistry In ASS Effected Ponds collected every two weeks to test the chemical variability over time. One of the main flaws in the study was the pH of the spot samples was not measured until returning to the laboratory. Variables such as pH need to be measured in-situ, as a temperature change will have a corresponding effect on pH. Also, the continuous monitors were calibrated every 1-3 months, increasing the possibility of instrumentation drift. Sensors were cleaned over the same time frames. This infrequent maintenance could present a problem in the data due to armouring (as iron could precipitate on the electrodes) or biofouling of the glass probes. Jarvie et al. (2001) corrected the continuous pH measurements using a calculated difference between the continuous measurements and the spot measurements. Data would have been of a higher quality if the monitors were calibrated and cleaned more regularly. Continuous monitoring is beneficial when monitoring extreme events that occur when the sampler is not there. However, the author of this thesis ensured that all physical water quality variables and reduced chemical elements were measured in-situ and analysed directly after obtaining a sample. This was to ensure that data was of the highest quality and a good representation of the chemistry at that point in the ponds.

4.3.1 Seawater Chemistry – Focus on Trace Metals

There have been a number of studies concentrating on trace metal accumulation in the water column of seawater. Krauskopf (1956) carried out experiments on the concentration of thirteen metals (Zn, Cu, Pb, Bi, Cd, Ni, Co, Hg, Ag, Cr, Mo, W, V), and one of the findings suggested that in an oxygen depleted environment, the concentration of eight of the metals (Zn, Cu, Pb, Hg, Ag, Cr, Bi, Cd) may be controlled by precipitating sulfides (which are associated to that particular local environment).

Byrne et al. (1988) performed calculations on over twenty metals in seawater. They found that strongly hydrolysed metals (such as Al3+, Cr3+, Fe3+, U4+ and Zr4+ etc) are greatly affected by pH and temperature, metals that are dominated by chloride complexation (Ag+, Au+, Cu+, and Cd2+ etc) and other metals (La, Ce, Gd, Yb, and Cu etc) generally complexed with carbonates.

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Fluctuations in the oceans metal concentration is related to depth in the water column. Hunter and Boyd (1999) noted that trace metal concentration in vertical depth profiles, follow a close relationship with the physical variables (such as dissolved oxygen, temperature and electrical conductivity). They commented that iron, generally found in seawater in its oxidised state of Fe(lll), is one of the most bioreactive trace metals in the ocean.

Analysis of trace metals in seawater has improved over the years due to improvement in instrumentation. Determining the actual species in seawater is still an area that needs a lot of work (Pesavento et al., 2001) and is generally limited by the large amount of Na+ and Cl- ions in the water. In the case of ICP- OES and ICP-MS (used in this thesis), water samples have to be diluted otherwise the high NaCl concentration block the injector and damage the machine. Most previous studies have concentrated on reporting trace metals as their "total" concentration in the sample. Metals are generally more toxic in their reduced form. Therefore in future studies (and as speciation analysis improves) it will be important to determine the oxidised and reduced fractions of a heavy metal and not just the total concentration.

4.3.1.1 Aluminium

Even though Al is the most abundant metallic element in the earth’s crust, it is only in very low concentrations in the ocean. Aluminium however is found in large concentrations in sediments associated with ASS or Acid Mine Drainage (AMD) where clays and Al minerals are dissolved in the acidic environment. Any free aluminium in the ocean is scavenged by biogenous material (e.g. siliceous particles in diatom blooms) and not recycled (Zhang et al., 1999).

Gibbsite is the most common Al mineral, with Al hydroxide minerals such as strandite and bayerite being less common. The size of the Al ion allows it to fit into a four-fold coordination with oxygen and allows it to substitute for silicon in tetrahedral structural sites. It can also form six coordination’s and can occupy octahedral crystal sites similar to those occupied by magnesium and iron. When the pH is acidic, the Al may be precipitated as an aluminium hydroxyl- sulfate. Nordstrom (1982) suggested that in acidic environments, the amount of Al dissolved in the water is controlled by the solubilities of gibbsite and

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Chapter 4: Review: Water Chemistry In ASS Effected Ponds kaolinite. Since most acidic waters are acid because of the oxidation of pyrite or sulfur, the dominant anion is sulfate. The aqueous geochemistry of Al is altered when there is sulfate present in the water. Generally, when the pH is less than 5, the cation Al3+ is dominant. In this form, it is likely that there are six water molecules surrounding an aluminium ion which is in the centre. It is possible that one of the water molecules may become an OH- ion if the pH increases slightly and this is when you get the formation of gibbsite.

Aluminium is insoluble at pH values above about 5.5, but is increasingly soluble at lower pH values (Dent, 1986). van Breemen (1973;1976) and Golez (1995) found that Al3+ activity is inversely related to pH, increasing roughly 10- fold per unit pH decrease. When Al is in its dissolved form, it becomes bioavailable and potentially harmful to plants and animals. Soluble Al can accumulate in the root tissues in plants, preventing cell division and elongation and, possibly, inhibiting enzymes concerned with synthesis of cell-wall material. The result is a stunted and deformed root system. In addition, uptake of phosphate is inhibited because of its absorption by aluminium in the soil and within the roots. In ASS’s, Al is the principle exchangeable cation. This is why exchangeable Al3+ is adsorbed and desorbed from the soil and the degree of adsorption is stronger in very acid conditions.

In aquatic environments, pH is an important factor contributing to the amount of dissolved Al in water. Broshears et al. (1996) looked at the behaviour of Al in acidic streams in America. They found that at a pH of 4.2, the Al concentration remained constant, however, when the pH reached 5.0 (an increase in pH was due to an upstream injection of sodium carbonate); there was a decrease in the concentration of dissolved aluminium and a corresponding increase in concentration of particulate aluminium. During the same experiment, Fe was observed to precipitate at pH 3.5. At pH 4.2, there was a slight decline in the amount of dissolved iron and an increase in particulate iron in the stream water and when the pH reached 5.8, dissolved iron decreased in size and particulate iron increased in size.

The toxicity of Al to aquatic organisms has been documented by a number of publications. Aluminium interferes with iron regulation in fish and also disrupts

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Chapter 4: Review: Water Chemistry In ASS Effected Ponds the important gas-exchange function of gills. Sammut et al. (1996) and Cook et al. (1999) found that Al caused most damage to fish at pH 5-5.2. In the latter study, water discharging from a cane farm at Pimpama (Qld) had a pH of 4.5 and still contained dissolved Al that exceeded the ANZECC Guidelines (2000).

The application of by-products from Al mining is becoming more popular for the remediation of ASS. Products such as Bauxsol™ (is the treated material that is left over after bauxite mining) have been neutralised with seawater and have been applied to ASS to decrease acidity due to its high acid-neutralising capacity (Lin et al., 2002). Bauxsol™ contains minerals such as hematite, boehmite, gibbsite, sodalite, quartz, cancrinite, brucite, calcite, diaspore, ferrihydrite, gypsum, hydrocalumite, hydrotalcite, p-aluminohydrocalcite, portlandite, minor aragonite and a few other low solubility trace minerals (McConchie and Clark, 2000). Virotec used Bauxsol™ at Tomei the year after this study as a trail (Final Report by Tomei Australia Pty Ltd). They found that there was a slight increase in the productivity in ponds where the product was used compared to previous years. However, Pagano et al., (2002) studied the by-products termed “red sludge” from a bauxite mine in Turkey. The sludge was coupled with sea urchin embryos (used as bioassays) to test if the bauxite by-products were having any detrimental affects. The sludge contained large concentrations of Al, Fe and other metals (Pb, Mn, Zn) depending on the bauxite composition. In one group of sea urchin embryos studied, there was 100% mortality rate when exposed to the “red mud”. They attributed the mortality to the pH being about 12 (at which Al has a high solubility and is in the form Al(OH)4). When the pH is alkaline (or at least about neutral), the - dominant dissolved form of Al is Al(OH)4 . When fluoride is present, strong complexes of aluminium and fluoride are formed (AlF2+). If the concentration of Fl is high the solubility of Al is increased.

High concentrations of Al in sampled water may be due to colloidal particulates passing through the 0.45μm filter paper during filtration (Kennedy and Zellweger, 1974; Hem, 1989). When there is a source of sulfate in the water, aluminium potassium sulfate or alum may form. This is generally used in water treatment processes to flocculate suspended particles and may leave a residue of aluminium in the resulting water (probably as small, suspended

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Chapter 4: Review: Water Chemistry In ASS Effected Ponds hydroxide particles). In aquaculture ponds, Al released from the soil matrix into receiving waters helps flocculate unconsolidated clays (White et al., 1997).

4.3.2 Stratification and the Sediment-Water Interface

Particulate organic matter found at the ocean surface undergoes rapid decomposition by bacteria while it sinks deeper in the water column towards the ocean floor. The particulate therefore has a different chemical composition to that when at the surface. It is hypothesised by the author that decomposing organic matter in prawn ponds (such as algae and added nutrients) also goes through a chemical change while sinking to the bottom of the pond. The distance the material travels isn’t as great as the distance in the ocean; however, it is affected by the mechanical motion provided by the paddle wheels and aerators in the ponds. These provide the means for the organic matter to mix with oxygen and trace metals (Al, Fe which can adsorb to the algae).

Aerobic mineralisation further degrades the organic matter at the sediment surface. Palmer et al. (2000) defined the sediment-water interface as the area where the water column comes into contact with the water saturated sediments. Smith et al. (2000) and Avnimelech and Ritvo (2003) stated that chemical flow across the sediment-water interface is generally in the form of particulate organic carbon and dissolved electron acceptors (e.g. oxygen, nitrate, iron, manganese, sulfate). These flow downward across the sediment- water interface and carbon dioxide, dissolved organic carbon, and various nutrients (e.g. ammonium, phosphate) are released from the sediments into the water column in marine environments.

Bioturbation of the sediment allows anaerobic bacteria to assimilate organic matter and during this process, releases ions to the pore water. This suggests why there is a high concentration of dissolved reduced ions in the sediment pore water. It also demonstrates how important the sediment-water interface is to chemical recycling of both fresh and marine environments (Jorgensen, 1983; Donahoe and Liu, 1998). A typical vertical sequence of reducing conditions occurs in the sediment. These chemical reactions are mitigated by denitrifying, sulfate reducing, and methane producing bacteria. These bacteria

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Chapter 4: Review: Water Chemistry In ASS Effected Ponds oxidise organic carbon to CO2, and ions such as N2, H2S and CH4 are produced (Jorgensen, 1983). Figure 17 depicts an idealised version of the distribution of oxidants and their relative distribution in the pore water of marine sediments.

Figure 17: Conceptual model for anaerobic decomposition (biologically mediated reactions) of organic matter, showing the exchanges between the atmosphere, water column and sediments (modified from Jorgensen, 1983)

4.4 Summary

Seawater naturally contains trace and heavy metals, however the oxidation of ASS releases a large concentration of metals contained in the sediment to the receiving water. Complex chemical reactions, occurring in the sediment pore water, assist in dissolving metals associated with minerals deposited in the sediment. Bioturbation and microorganisms facilitate the movement of these trace metals into the water column of the water body through the sediment- water interface. These trace metals (particularly Fe and Al) interfere with normal biogeochemical functions of aquatic organisms living in the water body. This may lead to the mortality of aquatic organisms. At Tomei, metals that are released from the sediments are discharged into the adjacent estuary (the process by which this occurs will be discussed in detail later in this thesis).

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5 REVIEW: AQUACULTURE - SPECIES, WATER QUALITY AND POND MANAGEMENT

5.1 Introduction

The types of chemical reactions that take place in the water column and at the sediment-water interface depend on the salinity and chemical composition of the water. Freshwater environments do not have the buffering capacity of seawater environments, and therefore when subject to ASS measured pH can be below 1. Acid sulfate soil affected aquaculture ponds containing seawater are naturally pH buffered and can have a water column pH that is typically between 6.97-8.99 (see data in Chapter 9).

5.2 Farming Regimes

Aquaculture species (such as prawns) are grown in freshwater or seawater environments using a variety of farming techniques. Based on the technique used, these environments are classified as being extensive, semi-intensive or intensive. Extensive aquaculture is defined as being when the natural environment supplies the prawns with food (Kautsky et al., 2000) and so typically requires there to be a low stocking rate in the pond. Juveniles are derived from wild populations, entering the ponds with the inlet water (Gräslund and Bengtsson, 2001). This is generally low maintenance farming; chemicals are not used (except fertilisers which may be used to promote the growth of algae as the food source). Extensive farming typically uses tanks and is commonly undertaken in estuaries where there is no anthropological input.

Semi-intensive aquaculture is defined as being when a proportion of the feed is supplemented by nutritionally formulated feed pellets, or fertilisers such as manure (Kautsky et al., 2000). Semi-intensive systems are set up using medium-sized ponds, and moderate stocking rates (about 10 prawns/m2). Juvenile prawns can be from wild stocks or bought from hatcheries. Food is either algae from the local environment, commercial pellets, or a combination of both (Gräslund and Bengtsson, 2001).

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Intensive aquaculture refers to managed ponds where the aquaculture species are fed with commercial food (Kautsky et al., 2000). The ponds are usually small, but have high stocking rates. Pond productivity relies heavily on artificial feed and mechanical aeration. The highly intensive nature of this method creates a high risk of disease from bacteria, viruses, fungi and other pathogens. This is due to overcrowding and stress experienced by the prawns (Gräslund and Bengtsson, 2001).

Both semi-intensive and intensive farming techniques use chemicals for medicating outbreaks of disease; cleaning and disinfecting ponds; growth of algae; eliminating undesirable aquatic organisms; and it is also common to add dye to the water to protect prawns against solar radiation during periods of decreased algal growth.

Approximately 72% - 80% of the world’s farmed prawns are farmed from the Asia-Pacific region (Trott and Alongi, 2000; Wolanski et al., 2000). Aquaculture is popular, particularly in poorer countries where healthy profits can be made in a short period of time (Sansanayuth et al., 1996). In Asia, prawn farms typically have a five to ten year life span. After this time, farms become unproductive due to disease and the associated mortality of the prawns: this rapid deterioration is largely due to underlying ASS. In these cases, farmers abandon the farms and re-establish elsewhere.

The Australian aquaculture industry is relatively small, it contributed a gross value of $733 million during 2001/2002, $64.4 million (9%) of that was attributed to prawn production (www.abs.gov.au/ausstats/[email protected]).

5.3 Prawn Characteristics

5.3.1 Prawn Species

There are three prawn species that are commonly farmed in Australia. They are: (1) the black tiger (or leader) prawn (Penaeus monodon); (2) the banana prawn (Penaeus merguiensis); and (3) Kuruma prawn (Penaeus japonicus) (Lobegeiger et al., 2001).

During the 2001-2002 season, Tomei farm management grew Penaeus japonicus in both Pond 7 and 10 (Plate 15). At the end of the grow-out season

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management these were harvested for sale live to the Japanese market; Pond 10 provided acceptable yields but Pond 7 provided very low yields.

Plate 15: Penaeus japonicus is the main species cultured at the Tomei prawn farm

Penaeus japonicus (Bate 1888) originated in Japan and are sought-after by the Japanese market primarily because of their brightly coloured tails. They earn about $15 million per year for the Australian export market (Burford and Glibert, 1999) and grow in subtropical regions. At Tomei these are farmed intensively and as such are fed artificial, pelleted, feed. In pond environments, Penaeus japonicus prawns have been observed to generally spend most of the day light hours buried in the pond bottom sediment (Egusa, 1961; McNeil, 2001). In the natural environment, their life cycle is rather different. Benzie (2000) studied the genetic diversity of naturally occurring penaeid prawns, of particular interest to this thesis are the two types of prawns grown at Tomei: P. monodon and P. japonicus (note: only P. japonicus were grown in the two ponds that form the basis of this study). In their natural environment their life cycle would be as follows (Table 5):

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Table 5: Information of species, geographic range and life history (adapted from Benzie, 2000)

Species Geographical Range Life History

P. monodon Indo-West Pacific (South Africa to India, SE Asia, Type 2a Fabricus 1798 Japan, north and east Australia) P. japonicus Bate Indo-West Pacific (South Africa to India, SE Asia, Type 3b 1888 Japan, north and east Australia) a Type 2 – adults spend their life at sea. They move into coastal waters to spawn, and the resulting larvae migrate inshore, usually into estuaries. The prawns stay in the estuary for a significant amount of time but move offshore to mature. b Type 3 – spends all its adult life offshore and (although the larvae are able to move into estuaries) they tend to inhabit inshore regions, living in sea grass or algal beds.

5.3.2 Prawn Stocking and Grow-out

Commercial farmers generally obtain their infantile prawns or "postlarvae" from hatcheries (e.g. Gold Coast Marine Aquaculture and Seafarm Australia). However, some farms (including Tomei) produce their own postlarvae (PL). This is done to minimise the potential of introducing disease to the farm. The postlarvae are stocked into (on average) 1 ha ponds when they have attained sufficient size (when the postlarvae are about 15 days old after metamorphosis from megalopa) and the pond water temperature is optimal for the species that is being introduced.

Stocking densities and grow out duration differs from farm to farm, however, commonly the Australian prawn aquaculture grow-out season for black tiger prawns is about 150-200 days. Stocking densities are generally about 350 000 - 600 000 PL’s per pond (35-60 PL’s/ m2), with prawns averaging 15-20 g at the end of the harvest.

In Asia and the Americas, the dominant species of prawn that is cultured is the Litopenaeus vannamei. This species is suited to be stocked at much higher rates up to approximately 700 PL’s/ m2. The grow-out season is much shorter averaging <100 days and prawns averaging between 15-35 g upon harvest. Again, these stocking densities, grow-out periods and harvest size are farm specific.

Six factors potentially limit the yield of prawns in an aquaculture pond: (1) little or no primary production; (2) high rates of pelagic respiration; (3) high suspended sediment loads; (4) slow rates of benthic decomposition; (5) low

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management and fluctuating pH and DO concentrations; (6) fluctuating salinity (Alongi et al., 2000).

5.3.3 Prawn Feed Requirements

In the natural environment, native prawns feed on sea grasses, copepods, gastropods, diatoms and algae (Loneragan et al., 1997). Prawns grown using either the semi-intensive or intensive methods are fed pellets (nutrient- enriched to supplement the prawn’s diet). These prawns also consume algae that live in the pond water column.

During certain stages of growth, prawns dietary requirements for healthy growth change. Juvenile prawns need a mixed diet of pellets and fresh food (such as mussel or squid). Without this fresh food supplement, they are likely to become nutrient deficient; this results in the development of white areas beneath the prawn’s exoskeleton (Sedgwick, 1979). Sedgwick (1979) demonstrated that when the prawns that developed symptoms of nutrient deficiency were fed fresh whole mussel, they recovered within two days. In Sedgwick’s 1979 study, the fresh mussel provided a better source of energy (i.e. fatty acids) than food pellets.

Fatty acids are only one component of the nutritional requirements for penaeid prawns. Proteins, amino acids, carbohydrates, lipids, vitamins and some minerals are important to a prawn’s diet as they are used for tissue repair and growth (Shiau, 1998). The required concentration of each of these is determined by which species was being farmed.

Penaeus japonicus are opportunistic carnivores (Reymond and Lagardere, 1990). Reymond and Lagardere (1990) found that even though prawns growing in semi-extensive ponds were being regularly fed pellets; they also consumed zooplankton, harpacticoids, chronomids and macrobenthic fauna.

Experiments have been carried out to replace feed pellets with cheaper alternatives. For example, cattle manure added to ponds increases the concentration of nutrients encouraging diatom communities to proliferate. Diatoms promote growth by supplying polyunsaturated fatty acids (Garson et

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management al., 1986; Wyban et al., 1987). Garson et al. (1986) found that food pellets are cost effective as they provide a greater prawn yield.

There is danger of overfeeding aquaculture species. Uneaten food and excreta are the main sources of nutrient waste in aquaculture ponds (Cripps and Bergheim, 2000). High nutrient concentrations in the pond can lead to pond eutrophication through the depletion of oxygen. Allan and Maguire (1995) undertook a controlled experiment in fibreglass tanks containing a deep layer of sediment; with organic matter (0.6%) and P. monodon (15 prawns/m2) placed in each tank. They varied feeding rates between the tanks and concluded that there was a decrease in the prawns’ tendency to graze over time. This was due to the pond sediment progressively becoming depleted in oxygen (finally becoming anoxic). The oxygen depletion resulted from the secondary metabolism of nutrients by other organisms in the tanks. They noted that the prawns were preferentially feeding from trays suggesting that the prawns were adjusting their habitats to escape the anoxic sediments. The authors concluded that water quality was poorer in the ponds with higher feeding rates (5.0% wet prawn biomass/day compared to 2.5% wet prawn biomass/day).

Development of anoxic pond water conditions results in the production of toxic metabolites, and is usually fatal for cultured organisms. The rate of oxygen consumption can be managed by improving: (1) the feed pellet integrity (ensuring slower breakdown rates); and, (2) optimising feed rates (to reduce the amount of uneaten food). This will reduce stress on culture organisms (brought on by hypoxia); reduce the amount of wasted food; and, thus improve the food consumption ratio for the pond(s).

5.4 Prawn Habitat Preferences

5.4.1 Penaeus japonicus – benthic inhabitants

Penaeus japonicus is a burrowing prawn and spends most of its resting time in the sediment (Egusa, 1961). They do however emerge from the sediment during feeding. McNeil (2001) observed juvenile prawns in the wild. He noted that the behaviour of prawns was different at night and during the day with the most notable changes at sunrise and sunset. During the daylight hours, the

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management prawns moved along the bottom in a tight pack (up to 250 prawns per square metre) and as the light intensity increased, the prawns packed closer together. If threatened, they scattered. If they came upon an area of protection, mangrove roots or substrate vegetation, they would disperse and hide in the vegetation. McNeil (2001) noted that wild adult prawns are harder to observe as they are less accessible (due to their life cycle taking them out to sea).

The P. japonicus and P. monodon are described by McNeil (2001) as being "benthically-orientated". These species are territorial and hunt for polychaetes in the bottom sediments. Other species marched around the pond in a teardrop shape (with the bulbous bit at the front) and when they sensed food pellets on the bottom of the pond, they would move towards them and pick them up while still moving. Accordingly many pellets were pushed into the sediment. P. monodon was observed to behave differently: they dig into the sediment: this results in pellets being broken into small fragments.

Uneaten food, prawn faeces, dead and decomposing aquatic organisms, soil particles all constitute waste on the bottom of prawn ponds (Boyd 1992; Paez- Osuna et al., 1999). Water quality in aquaculture ponds deteriorates when the amount of organic matter or introduced nutrient is excessive. The presence of this excess material causes ammonia, organic sulfur, and in ASS environments, hydrogen sulfide to be released into the pond water (Funge- Smith and Briggs 1998).

Penaeus japonicus (being burrowing prawns) spend most of their time at or below the sediment-water interface. They are therefore sensitive to chemical and physical conditions and changes at this interface. The sediment can be either a source or sink of nutrients, trace and heavy metals as it continuously interacts with the water column through the sediment-water interface (Bratvold and Browdy, 2001).

Peterson (1999) stated that benthic shear stress (mechanical erosion) occurs when there is water movement in the pond. Shear stress enables substances (such as oxygen) to diffuse through the sediment-water interface. Similarly this reaction to water movement can lead to the localised concentration of organic matter, nutrients or waste. The preferential accumulation in certain pond areas

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management can lead to nutrient-enriched areas on the pond floor. The bacteria associated with decomposing organic matter and other nutrients will consume oxygen faster than it is able to diffuse from the overlying oxic water column. This leads to the establishment of localised anaerobic conditions in the sediment.

Pond mechanical aeration is designed to combat this problem; however, it induces pond floor and dyke wall scouring, sediment transport and accumulation. The pond floor sediment is also affected by: the amount of prawns stocked in a pond; the feed; the type of aeration used; the number of installed devices; and, the material from which the pond dyke walls are constructed. These variables can affect the type of conditions at the sediment- water interface by either mechanical or chemical interaction.

5.4.2 Aerators

There are two main types of aerators used in aquaculture, they are (1) diffusion aerators (Plate 16) and (2) spray-type surface aerators (such as the paddle wheels – Plate 17) (Viso, pers com. 2004). Diffusion aerators direct air towards the bottom of the pond and are important in relatively deep ponds (or ponds where bottom dwelling aquaculture species are being grown). The spray-type surface aerators are generally used to mix the water column (Boyd and Martinson, 1984).

Plate 16: Aspirators at Tomei

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Plate 17: Paddlewheel aerators at Tomei

Aerators are designed to deliver oxygen to the water column and mix the pond water to prevent either thermal or chemical stratification. However, their effectiveness is questionable. Previous studies have concentrated on the efficiency of aerators in fish ponds (Tucker and Steeby, 1995; Boyd, 1998), and since fish live in the water column, paddle wheels and oxygen diffusers supply an adequate concentration of DO. Tucker and Steeby (1995) found that in shallow (1m deep) catfish ponds that aerators led to water circulation which increased the DO at the sediment-water interface. In prawn ponds the farmed organisms spend most of their time in the benthic zone. Aerators have been shown through this study to provide insufficient quantities of oxygen. The configuration of aerators should be chosen based on: (1) the aquaculture species being grown in the pond; (2) local climate and, (3) size of the pond. This study shows that ponds become stratified and that DO concentrations at the sediment-water interface are typically inadequate for benthic aquaculture species.

Farm management practices at Tomei led to the aerators being turned off during the evening (Viso, pers. com. 2001). Periods of darkness are when pond DO concentrations decrease. During the day photosynthesising algae provide oxygen to the water column; at night, respiration by pond organisms (including algae) depletes oxygen. Accordingly, the DO concentration in the water column (and therefore at the sediment-water interface) becomes

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management dangerously low for the aquaculture species: leading to stress and reduced productivity or potentially mortality.

Aerators can be useful in delivering oxygen to the water column; however they may cause erosion of the pond bottom, and the dyke walls (Boyd, 1998). To avoid eroding pond dyke walls, pond aerators and paddle wheels should be placed in a configuration that points towards the pond centre (Avnimelech and Ritvo, 2003).

In contrast, Delgado et al. (2003) stated that it is typical during intensive aquaculture practices for the aerators to be placed parallel to the pond dyke walls, and the centre of the pond receives less water flow (aeration) than the edges of the pond. These authors trialled this approach and showed that there are two main areas in the pond: (1) the well mixed oxygenated area located near the dyke walls; and (2) relatively stagnant, stratified region in the centre of the pond. The central area was found to contain a larger proportion of accumulated waste and organic material.

5.4.3 Dissolved Oxygen

Dissolved oxygen enters the pond water in the following ways: (1) mechanically with the use of aerators; (2) from photosynthesising algae during the day; (3) from atmospheric sources entering the pond by diffusion; and, (4) through external water exchanges.

Dissolved oxygen is depleted from pond water by: (1) respiring organisms (both plants and animals); and, (2) biochemical and chemical reactions at the sediment-water interface. The DO concentration can also decrease when the pond temperature is higher, or when the salinity or barometric pressure changes (Boyd, 1998).

Rosas et al. (1999) indicated that temperature, salinity, pH, pollutants, activity and size are some factors that determined changes in the metabolism of animals exposed to hypoxia (decreased DO), and that penaeid prawns were particularly sensitive to low concentrations of DO.

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Oxygen is produced during the day by photosynthesising algae and several studies conclude that photosynthesising algae are the most important natural source of oxygen in ponds (Chang and Ouyang, 1988: Boyd and Teichert- Coddington, 1992; Boyd, 1998). Schroeder (1975) stated that when ponds stocking levels are correct, the photosynthetic production of oxygen by algae is sufficient to meet the total oxygen demand within the pond.

Photosynthesis occurs in response to sun light and therefore, occurs most frequently at the top of the water column (where light intensity is the greatest). It is also here that direct diffusion of oxygen from the atmosphere occurs, and thus the top of the water column is generally saturated with DO. However, as either the culture species stocking density or the rate of organic fertiliser addition is increased, pond oxygen demand increases. Pond oxygen demand peaks in response to: (1) respiration of the culture species, plankton and bacteria active in the decomposition of fish excretion; (2) manure and uneaten portions of the supplied food.

Culberson and Piedrahita (1996) identified the DO profile through a water column drops during night hours. They stated that previous studies assumed that the pond water column is uniform over depth. However, they found that shallow ponds (without aeration devices) tended to stratify during the day with respect to DO and temperature, and mix during the night (Chang and Ouyang 1988, Losordo 1988).

Typically the greater the water depth, the less DO is present in the water column. Meijer and Avnimelech (1999) measured oxygen in undisturbed pond sediment with a small tip oxygen probe. They observed a drop in oxygen concentrations from fully saturated to 50% saturated in a 1mm interval of the water column (just below the sediment-water interface). At 1-1.25mm, no oxygen was measured. They suggested oxygen diffusion into the sediment was facilitated by eddy flow and stated that the force of viscosity near the sediment-water interface allowed the water to "stick" to the surface of the sediment. They named this layer the diffusive boundary layer (DBL) and suggested that within this layer, molecular diffusion of solutes occurs. It is also important to recognise that oxygen can be introduced into anaerobic zones by

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management aquatic animals and plants. These biota create burrows in sediments which increases surface area allowing more oxygen to be diffused into the sediment (Kristensen, 2000).

In aquaculture ponds, farmers attempt to ensure that oxygen is mixed into the deeper sections of the pond by the mechanical action of aerators and/or paddle wheels. In areas exposed to wind this also assists atmospheric DO diffusion. However the bottom of the ponds is seldom saturated with oxygen due to the inefficiencies in circulation and mixing. In the pond sediment microbiological communities (such as heterotrophic microbes and bacteria) consume DO while feeding (Bratvold and Browdy, 1998). In addition to being fatal to prawns, a lack of oxygen at the bottom of the pond causes the toxicity + of NH4 (ammonium) to increase. A rapid decline in DO is often linked to an increase in the concentration of ammonia (NH3) and decrease in pH (Chang, 1986).

Table 6 provides a list of mechanisms and their effects for increasing the concentration of DO in the water column of aquaculture ponds.

Table 6: Mechanisms for increasing the concentration of DO in aquaculture ponds (adapted from Boyd, 1998).

Mechanisms to improve DO in pond water Effect of mechanism Combining the use of aerators and fertiliser increase oxygen production by photosynthesis of aquatic plants Adding chemicals to pond water promote the release of oxygen Aerating ponds with pure oxygen gas Delivering a greater concentration of DO to pond water Mechanical agitators oxidise the water whilst exposed to the atmosphere Use of air stones Oxidises water in-situ

Numerous studies have concluded that during a normal daily cycle, the concentration of DO is the lowest just before dawn (Chang, 1986; Madenjian et al., 1987; Tucker and Steeby, 1995; Boyd, 1998). This phenomenon is attributed to dawn being the longest period of time since the photosynthesising organisms were actively producing oxygen. Boyd (1998) measured the DO concentration at sunrise to be as low as < 2 mg/L. Concentrations below this can lead to stress and mortality of the aquaculture species. Garson et al.

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(1986) notes that when prawns come to the pond surface (in the early hours of the morning due to low DO), they become feed for predators such as birds.

Depleted DO and prawn mortality have been linked in numerous studies (Egusa, 1961; Allan et al., 1990; Allan et al., 1995). The result of DO depletion varies with the prawn species. Lethal concentrations of DO for P. monodon (juvenile) is about 0.9 mg/L (Allan and Maguire, 1991); P. monodon (adult) is 0.1 mg/L (Allan et al., 1995); and, between 0.72 mg/L and 1.43 mg/L for P. japonicus (whether resting or active) (Egusa, 1961).

Extremely low DO concentrations can reduce growth, feeding and moulting frequency (Allan and Maguire, 1991). One of the causes for rapid fall in pond DO is the decomposition of microbes on the pond base after a rapid algal die- off. This process also alters other water quality variables (such as ammonia). Allan et al. (1990) found that DO concentrations of around 2.1 mg/L, enhanced the acute toxicity of ammonia to P. monodon. He noted that the prawns affected by low DO concentrations, initially swam to the surface. Allan et al. (1990) found that the amount of oxygen the prawn consumes was dependent its size; and noted that the extent of damage caused by low DO was dependent on the farm managers response to the problem.

Egusa (1961) ran an experiment observing P. japonicus prawns (3-17g) and their response to changes in DO. Egusa (1961) found that if the oxygen concentrations were optimal, the prawn would stay buried in the sand (primarily during the day); however, if oxygen concentrations fell to around 1.43 mg/L, the prawn would move into the water column. A summary of Egusa’s (1961) findings is contained in Table 7.

Table 7: Concentration of oxygen and physical response from prawn

Measured concentration of Physical response from prawn oxygen (mg/L) 1.43 respiratory siphon pokes out of sand 1.14 prawns rostrum and eyes protrude from sand 1.14 1 0.72 cephalothorax protrudes from sediment 0.72 prawns to the surface of sand and start showing signs of suffocation <0.72 prawns suffer mortality

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In Egusa’s 1961 experiment, oxygen was reintroduced into the tank once the prawns came out of the sand. They were observed to bury themselves again when oxygen concentrations reached 1.43 mg/L. It was noted that the smaller prawns were more sensitive to a lack of oxygen than larger prawns.

5.4.4 Nutrients

There are two dominant inorganic forms of nutrients associated with prawn - + 3- aquaculture ponds: nitrogen (NO2, NO3 and NH4 ) and phosphorous (PO4 ).

Major nutrient gain to pond water is through: water inflow (from water exchange with a nutrient-rich estuary, rainfall and runoff); fertilisers (used to promote algal blooms); the breakdown of excess feed; aquatic animal excretion; and, nitrogen released from decomposing organic matter. Cyanobacterial nitrogen fixation and atmospheric deposition are an occasional source of nitrogen (Hargreaves, 1998).

5.4.4.1 Nitrogen

Only 20%-40% of the nitrogen in pelletised food is absorbed into prawn tissue, the remainder is lost as excreta to the pond (Lorenzen, 1999). This causes nutrients to be discharged from ponds during water exchanges which impacts the adjacent environments. If the waste/excess nitrogen is not discharged from the pond during water changes, it can accumulate in the pond sediments and is subsequently released once the pond is drained at the end of the grow-out season. Lorenzen (1999) suggests three methods for recovering nitrogen from pond effluent: (1) aquatic plants uptake of dissolved N; (2) uptake of N by filter feeding organisms (mainly phytoplankton); and, (3) uptake of particulate and dissolved N by plant beds. An integrated approach utilising methods is likely to be the most effective means of removing N-rich aquaculture effluent.

Aquaculture ponds receive estuary water during water exchanges. Anthropogenic sources can cause intake water to be nutrient-rich. Increased levels of nutrients cause eutrophication of the water and vigorous growth of phytoplankton and macroalgae. Often in this situation, aquatic plants (located at the base of the water column), suffer due to the lack of light penetration through the water column. This results in a change in the location of

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management photosynthesising organisms, increases oxygen consumption and results in lower DO concentrations in the water column. This process also increases the oxygen consumption at the sediment-water interface due the deposition and rapid degradation of the associated biomass (Herbert, 1999). The + - accumulation of toxic inorganic nitrogenous species (NH4 and NO2 ) is a key reason for poor water quality. This is best addressed through the introduction of oxygen to the pond: ponds that are highly aerated contain enough DO for the bacteria to oxidise ammonium to nitrite and nitrate (Avnimelech, 1999).

+ Most algae and microorganisms prefer ammonium (NH4 ) as the inorganic nitrogen source (Avnimelech and Zohar, 1986). In this form the nitrogen is in the reduced (-III) oxidation state making it suitable for cell synthesis. When ammonium is not available, many prokaryotic cells use the oxygenated forms - - of nitrogen (such as nitrate (NO3 ), nitrite (NO2 ) and di-nitrogen (N2)). When these forms are used, microorganisms must initially reduce the oxidised N- form to the (-III) reduced oxidation state. This process requires energy and electrons.

Inorganic nitrogen is only one element that affects the nitrogen fixating capacity of bacteria. The availability of carbon, trace metals, light, dissolved oxygen, a suitable pH, correct temperature, and salinity also influence their metabolic rate. There are four steps that control the nitrogen cycle in a marine environment. They are (1) ammonification; (2) nitrification; (3) denitrification and, (4) nitrate ammonification. Each step is briefly described below: (1) ammonification – ammonium is released from sediment or organic matter; (2) nitrification – oxidation of ammonium. It is a two stage process. Ammonium oxidising bacteria produce nitrite, which nitrite oxidisers further oxidise to nitrate; (3) denitrification – heterotrophic bacteria use nitrate as a terminal electron-acceptor during respiration and reduce nitrate to N2 gas (denitrification) or ammonium (nitrate ammonification) (Herbert, 1999).

Ammonia is released from the system by phytoplankton uptake and nitrification. The nitrogen biochemistry of an aquaculture pond is affected by: (1) the amount of N contained in feed and the frequency of feeding; (2) water exchange and circulation; (3) aeration; (4) pond depth and other management

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management procedures. Ammonia is very toxic to aquatic animals because it competitively binds nitrite to hemocyanin in crustaceans and haemoglobin in fish. This changes to methemoglobin (which does not have the carrying capacity of oxygen) which in the presence of nitrite causes hypoxia and cycanosis (Chen and Cheng, 2000). Fish with a higher percentage of methemoglobin suffer from functional anaemia due to the reduction of oxygen-carrying capacity. It can also cause damage to the cerebral energy metabolism and to the gills, liver, kidney, spleen and thyroid tissue in fish, crustaceans and molluscs (Allan et al., 1990). Ammonia has a higher toxicity to aquatic organisms when the pH and temperature is elevated (because the ionisation equilibrium moves towards the toxic, unionised gaseous form).

Ammonia flux from sediment can be enhanced by bioturbation. Benthic invertebrates burrow into the sediment and increase the flux of ammonia into the water column by up to 50%. Some burrows extend 80-120 mm into the sediment. The accumulation of nitrate in pond sediments is caused by denitrifying bacteria consuming nitrate and reducing it to N2 below the oxic zone. The sediments can act as either a source or sink for nitrate to the pond water. The reduction of sulfate to sulfide minerals or gas in the sediments may inhibit nitrate reduction by consuming the available oxygen and reduce the accumulation of N2O and NO (Jorgensen, 1983).

Prawns also influence the nitrogen cycle by burrowing into the sediment and by producing faecal pellets. Herbert (1999) noted that faecal pellets are sites of intense microbial activity which results in a high oxygen demand. Microniches where sediments undergo active reduction allow denitrification to occur in otherwise oxic surface sediments.

+ Ammonium (NH4 ) may weakly adsorb to negatively charged cation-exchange sites on the surface of clay minerals or organic matter in the sediment. For nitrification (2 step oxidation of ammonia to nitrite) to occur there needs to be:

Nitrosomonas and Nitrobacter to catalyse the reaction, and 2 moles of O2 for + every mole of NH4 (Hargreaves, 1998).

Nitrate is produced in most sediment by the oxidation of ammonium which naturally diffuses upwards into the lower part of the oxidised sediments. In this

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management setting nitrate is an intermediate product in the nitrification of ammonia and a product in the denitrification of nitrate.

- In water, there are two forms of nitrite: (1) nitrite ion (NO2 ) and (2) nitrous acid

(HNO2). The latter form is freely diffusible across the gill membrane. Nitrite ion is not freely diffusible, which is fortunate because it is toxic to aquatic animals.

In a recirculating water system, the accumulation of nitrite and nitrate may result in a fall in pH (Chen and Cheng, 2000). P. japonicus and other species of prawns are sensitive to concentrations of nitrite higher than 100 μg/L NO2-N (Mevel and Chamroux 1981; Chen and Chin 1988). In environments that are affected by high temperatures and low oxygen concentrations, increased respiration leads to more water passing over the gills. This process also allows the diffusion of nitrous acid which can result in mortality (Chen and Chin 1988; Chen and Cheng, 2000).

Nitrogen is lost from the ponds through: (1) water outflow (drainage, seepage, and evaporation); (2) prawn harvest at the end of the grow-out cycle; (3) N loss to the atmosphere due to volatilisation of ammonia; (4) sediment burial; and, (5) denitrification.

Briggs and Funge-Smith (1994) undertook research that focused on nutrient budgets of intensive marine aquaculture ponds in Thailand. They concluded that the sediments are the major sink of nutrients (N about 30% and P about 84%) in intensive shrimp ponds. They suggest that between grow-out cycles it is extremely important to clean out the bottom of the pond. Cultured aquatic organisms use between 5-40% of nutrients supplied in feed. The average nutrient retention by prawns is 29% (of the total pond budget) for nitrogen and 16% for phosphorous (Avnimelech and Ritvo, 2003).

Chen and Chin (1988) found that P. monodon could not moult cleanly at low nitrite concentrations and found that by increasing the affinity of hemocyanin for O2 (at the time of high metabolic demand) it decreases the prawns’ resistance to nitrite toxicity during moulting.

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5.4.4.2 Phosphorus

Phosphorus promotes the growth of photosynthetic algae and cyanobacteria. It can be removed from a system by: (1) uptake into biomass; (2) precipitation. It can be added to a system such as a pond environment from: (1) fertilisers; and, (2) dissolution of solids.

Phosphate may precipitate as authigenic minerals by combining with iron, manganese or calcium (vivianite: Fe3(PO4)2.8H2O; or apatite:

Ca5(PO4)3(OH,F)) in reduced sediments. Under oxic conditions at the sediment-water interface phosphate is bound by adsorption to ferric oxyhydroxides. If the redox conditions at this interface change, a slug of phosphate may be released (Jorgensen, 1983).

Iron-rich sediments are often associated with ASS environments. Dent (1986) stated that ASS are associated with nutrient deficiencies, particularly deficiencies of phosphate. This may be related to the high concentration of active aluminium and iron, which form insoluble phosphates at low pH. Brinkman and Singh (1982) and Singh (1981) found that when phosphate is added to ASS affected pond water it is rapidly absorbed by aluminium salts or by unbound aluminium in the sediment. These reactions result in high Al and 3- low PO4 concentrations and inhibit the growth of the algae. Furthermore these authors noted that after heavy rain ferric hydroxide formed in the water column of the fish ponds and clogged the gills of the fish, killing those that survived the acid slug from the dyke walls.

Phosphorus concentrations have been noted to vary depending on which part of the pond the sample is taken from. Smith (1996) found that the pond walls and outer edges of the pond had the lowest concentrations of phosphorus, and that the highest concentrations were in the centre of the pond.

5.4.5 Water exchanges

Water exchanges are undertaken by farm managers in aquaculture ponds to: (1) increase the concentration of DO in the pond water; (2) decrease the salinity of the pond water; and, (3) discharge waste material (including a build up of nutrients and toxic metabolites such as ammonia).

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During water exchanges water is taken in from near-by estuaries through intake pipe(s) and is usually held in a settlement pond before it is discharged into the aquaculture ponds. The settlement ponds are used to remove suspended solids from the piped estuary water.

During intake, some of the water in the pond needs to be discharged before the “clean” water is exchanged. Water from ponds is discharged to the estuary from the ponds. Sometimes the water may contain high concentrations of: (1) nutrients; (2) metals; (3) pond organisms (which may or may not be infected with disease); and (4) pond sediment (such as ASS).

Pond water exchange is not the only mechanism of gaining or loosing water from ponds. In some ponds water seepage through the bottom or through the dyke walls occurs. Firm compaction of the pond bottom sediment and walls can reduce the likelihood of seepage occurring. In some cases the pond bottom can be sealed with clay, or seepage can be slowed by mixing organic material into the soil (Boyd and Gross, 2000).

5.4.6 Microorganisms

Herbert (1999) states that even though 79% of the earth’s atmosphere is made up of nitrogen, plants and animals are unable to make use of this supply without nitrogen fixating microorganisms. Cyanobacteria are the dominant form of nitrogen fixating bacteria in the marine environment, however, heterotrophic bacteria also fixate nitrogen, the genera and environments of these are presented in Table 8.

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Table 8: Heterotrophic bacteria that fixate nitrogen in marine environments (modified from Herbert, 1999).

Genus Habitat Anaerobes Sea grass sediment Azotobacter spp. Intertidal sediment Estuarine sediment Salt marsh sediment Microaerophiles Spartina roots Azospirillum spp. Zostera roots Spartina roots Campylobacter spp. Sediment Beggiatoa spp. Facultative anaerobes Beach sediment Enterobacter spp. Intertidal sediment Klebsiella spp. Estuarine sediment Seawater Vibrio spp. Zostera roots Anaerobes Marine sediments Desulfobacter spp. Sea grass sediment Desulfovibrio spp. Intertidal sediment Salt marsh sediment Estuarine sediment Sea grass sediment Clostridium spp. Intertidal sediment Salt marsh sediment Estuarine sediment Archaea Seawater Methanococcus spp. Estuarine sediment Methanosarcina spp.

Microorganisms are also useful in aquaculture as they: (1) ensure good water quality (regulate pH and ammonia concentrations); (2) regulate the oxygen content in the water; (3) are supplementary food sources to aquaculture species; (4) restrict disease and pathogens; and, (5) assist in the biochemical breakdown of accumulated wastes (Moriarty, 1997; Bratvold and Browdy, 1998).

The sediment-water interface is the most bioreactive area of an aquaculture pond. At the sediment-water interface, the sediment typically contains some oxygen from the water column. As this oxygen slowly diffuses into the sediments, microorganisms consume it as they assimilate organic matter and convert it to CO2. This process results in the sediment becoming anoxic and creates conditions which led to the proliferation of sulfur-reducing bacteria (Berner, 1984; Calmano et al., 1993). Deeper in the sediment, microorganisms

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management use oxygen from oxic compounds in the sediment; this decreases the Eh (redox potential) of the sediment. Under anoxic conditions, minerals such as pyrite are stable, however, if conditions change and they are exposed to oxygen (by microbial interaction, groundwater intervention, sediment excavation) these minerals oxidise and any associated heavy metals become mobile.

Meijer and Avnimelech (1999) found that there are large numbers of microbes in the surface sediment of fish ponds. These microbes use the available oxygen; inhibiting it from moving deeper into the sediment. An oxygen electrode measured the penetration of oxygen of up to a few millimetres in freshwater and marine lakes and 1mm in intensive and semi-intensive pond sediments under calm conditions. They measured oxygen consumption by the organisms contained in the sediments to be approximately 45-50mg 2 O2/m /hour at 25°C. By measuring and understanding the redox profiles, they found that sulfide formation occurred at 2-4mm below the sediment surface. Bioturbation, wind driven agitation, or water movement can increase the depth that oxygen penetrates the sediment-water interface, which alters redox conditions and can lead to chemical reaction.

Avnimelech and Ritvo (2003) stated that aerobic bacteria in the sediments of fish and prawn ponds use large quantities of oxygen during feeding and assist in depleting oxygen from the sediments, turning them anoxic. It is in these anoxic sediments that anaerobic bacteria thrive and in doing so they continue the chemical evolution of the pore water sediments as they grow by processing the available electron acceptors (other than oxygen) (Figure 18).

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Figure 18: Sequence of electron acceptors

Fertilisers are used during prawn production for promoting benthic organisms and phytoplankton growth in the pond. These organisms provide an alternative food source to the prawns, contribute to the pond oxygen cycle through photosynthesis and decomposition, assist in controlling pH, absorb toxic metabolites, and reduce the amount of light penetration to the bottom of the pond (reducing stress)(Briggs and Funge-Smith, 1994).

5.4.6.1 Sulfur Reducing Bacteria

Sulfur reducing bacteria are important to the chemistry of marine environments. Jorgensen (1983) stated that up to half of the oxidation of organic matter in the sediments of a marine shelf environment (the main site of mineralisation in the sea) is undertaken by sulfate reducers, whereas only 3% is processed by denitrifying bacteria. This demonstrates the oxidation of organic matter provides the link for cycling nutrients (that are essential for the growth of organisms such as phytoplankton) into the sea water. These microalgae are also important food sources for higher animals in the food chain.

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Sulfur reducing bacteria are mostly heterotrophic. That is, most of the carbon that they metabolise is derived from organic matter. Most strains of Desulfovibrio do not fix nitrogen, and those that are able to prefer ammonia as their nitrogen source. The efficiency of ammonia metabolism (in terms of cell growth and energy production) is twice that achieved during nitrogen fixation (Goldhaber and Kaplan, 1975).

Subtropical estuarine environments generally provide optimal conditions for sulfur-reducing bacteria to proliferate. These environments contain large amounts of organic matter and sulfur from seawater. The bacteria reduce the sulfate in the seawater to sulfide (in the form of hydrogen sulfide); organic matter to CO2 or hydrogen sulfide (H2S); and, ferric iron in the sediments to ferrous iron (detrital iron oxides with pyrite being the end product) (van Breemen, 1988; Moriarty, 1997; Wilson et al., 1999; Preda and Cox, 1998a).

Sulfate reducing bacteria are also able to metabolise (and therefore reduce): nitrite to ammonia; arsenate; disproportionate thiosulfate; or, elemental sulfur to sulfate with Mn(IV) as an electron acceptor. Other sulfate reducing bacteria can enzymatically reduce Fe(III), U(IV) and Cr(VI) but do not grow while these metals are used as the terminal electron acceptors (Coleman et al., 1993; Tebo and Obraztsova, 1998). These bacteria are able to survive in an oxygenated environment by remaining dormant (no growth occurring) until anaerobic conditions are reinstated (Le Gall and Xavier, 1996; Alongi et al., 1999; Cypionka, 2000).

Preda (1999) stated that a system needs to be free of oxygen for the sulfur- reducing bacteria (such as Desulfovibrio and Desulfotomaculum) to be present. However, Cypionka (2000) studied oxygen respiration by Desulfovibrio species and stated that sulfate-reducing bacteria are able to live in the presence of oxygen for several days and reduce oxygen to water.

In estuary systems sulfate-reducing bacteria live in the oxic and oxic-anoxic interfaces of sediments, in stratified water columns, microbial mats and in the gut of termites. Sulfate-reducing bacteria tend to clump together, or form aggregates with compounds such as iron sulfides, when they are exposed to oxygen. Sulfur-reducing bacteria that aggregate have a higher survival rate in

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management an oxic zone than those that remain free-living. Cypionka (2000) noted that even though sulfate-reducing bacteria acquire their name from one element, they are able to reduce many electron acceptors, not just sulfur. This is the case for the reduction of nitrate or nitrite to ammonia. During this process, the sulfate-reducing bacteria grow larger than they do when reducing sulfate. Growth of these species is not, however, linked with reduction of metal ions. Electron acceptors, arsenate (As(V)) and Mn(IV) are linked with the growth of Desulfotomaculum species. Some Desulfovibrio species are observed to stop growing when exposed to oxygen, but start growing again when returned to anoxic conditions. In sulfate reducing environments, such as monosulfidic soils or estuarine environments, the black layers are assumed the site of sulfate reduction. The colour of the sediment shows that the sulfur is in a stable form. This is because if sulfate reduction continues in the oxic region, the black sulfides would be reoxidised quickly and would not accumulate.

Boyd (1992) stated that the optimal temperature for microorganisms is in the range of 25-35°C. Due to the microbiological activity in aquaculture ponds, oxygen is generally used up faster than it can be delivered. This is particularly the case in ponds that contain large amounts of organic material.

5.4.7 Phytoplankton

Phytoplankton has a significant impact on the water quality of aquaculture ponds (Chien, 1992). It regulates the amount of DO in the water through its control of photosynthesis. Carbon dioxide, NH3, NO2, and H2S concentrations can decrease as oxygen is released into the pond.

+ Phytoplankton can consume NH4 and in doing so, binds heavy metals. However if the prawn stocking density is high (98 animals m-2) in the pond, the concentration of ammonia is too great for the phytoplankton to assimilate it (Burford and Glibert, 1999).

Phytoplankton aids in filtering light and therefore: (1) reduces the growth of filamentous algae on the pond bottom; and, (2) provides protection for prawns against predators. Phytoplankton also assists in stabilising pond water temperature, decreasing the amounts of pathogenic bacteria and provides an alternate food source for prawns. Vigorous photosynthesis by phytoplankton

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management blooms can cause pond pH to increase to above 8 and can also increase NH3 to concentrations that are toxic to aquatic organisms. The proliferation of phytoplankton can decrease the biomass of the zooplankton and bacteria (Arauzo et al., 2000). It is important to balance the three organisms to ensure a healthy pond.

Phytoplankton growth alters the colour of the pond. Chien (1992) states that green algae and phytoflagellates are normally more stable than the diatoms and zooplankton (which give a brown tinge). Once the phytoplankton reaches the end of its life cycle, or the water is no longer suitable for it to live, it dies off. During the decay of the phytoplankton, toxins may be released into the water which may harm prawns (increased waste organic material from decomposing cells and therefore a reduction in water quality), or the release of ammonia. If a mass of phytoplankton die, they are usually seen on the surface of the pond as a thick floating foam, and usually then sink to the bottom of the pond where decomposition occurs (Funge-Smith and Briggs 1998; Neori et al., 1989). During periods of phytoplankton crashes, ammonia-N tends to accumulate in the pond and can result in high measured nitrogen concentrations.

Mass moulting of prawns has been observed at the beginning of phytoplankton die off. Phytoplankton can deplete DO from the water on overcast days, or as they respire at night.

Water discharged into estuaries from aquaculture ponds contains effluent (in + the form of NH4 ) which promotes the proliferation of phytoplankton blooms in the estuary. The blooms can cause skin irritations, breathing difficulties and eye irritations in humans. Dermatitis has been linked with mats of Lyngbya majuscula (“Mermaid’s Hair”) which is a member of the cyanobacteria family. Proliferation of this species has been associated with high iron concentrations. The iron is used by the bacteria during nitrogen fixation and in Deception Bay, has been linked to ASS runoff (Dennison et al., 1999).

Studies carried out by Dennison et al. (1999) showed that the phytoplankton (in Moreton Bay, Australia) appeared to prefer nitrogen as a source of nutrients over phosphorus or silicate.

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5.4.8 Disease

Prawns grown in aquaculture ponds are susceptible to disease. Disease reported during prawn production is typically associated with stress due to inferior water quality or overcrowding (Davis and Arnold, 1998).

Pathogens that cause the disease may be introduced to the prawn in the hatchery, or may enter the pond through the exchange water. Diseases such as: “yellow head” (YBV); “white spot“ (SEMBV); monodon baculo virus; hepatopancreatic parvo; and infectious hypodermal and hematopoetic necrosis, can attack prawns, and can be carried by healthy prawns to infect others which can then cause mortality (Funge-Smith and Briggs, 1998; Gräslund and Bengtsson, 2001).

Disease causing pathogens can remain dormant in the prawn until it is stressed. Stress may be brought on by: (1) poor water quality; (2) repeated fluctuations of O2, salinity, and temperature; (3) high stocking density; (4) closely located ponds; and in ASS environments, (5) acidified soil releasing heavy metals from the sediments (Kautsky et al., 2000). Maintenance of water quality, reduced stocking densities and large buffer zones between ponds; and the implementation of farming practices that are in line with preservation of the environment aid in decreasing the frequency and impact of disease in aquaculture ponds.

5.4.9 Harvesting

At Australian prawn farms, the common first pass method for harvesting is with the use of nets or traps. Traps are placed at the bottom of the pond and are set with a bait (such as mussel) to encourage the prawns to enter the trap. Once the prawns are inside, they are unable to exit the trap. Traps are generally set the day before a harvest, to enable time for prawns to enter. Harvesting can extend over a few nights. When the density of prawns caught in the trap decreases, the pond is drain harvested. As the name suggests, ponds are drained of water and prawns buried in the sediment generally come to the surface. This is a more labour intensive process, as farm staff have to collect the prawns from the bottom of the ponds.

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5.4.10 Pond water temperature during grow-out and shipment

The optimal water temperature for P. japonicus is between 24-29°C (Coman et al., 2002). However, Hewitt and Duncan (2001) found that when pond temperature was between 28°C and 32°C prawns ate more, moulted more frequently, and had better survival rates. They found that high pond water temperature caused P. japonicus physiological stress. They also found that mortality was the greatest at 36°C, with 100% of the prawns dying after 12 days. They suggested three ways of ensuring optimal water temperature during the production cycle: (1) timing production to seasons with optimal temperatures; (2) increasing the pond depth; and, (3) increasing the frequency of water exchanges.

Prawns that are harvested for live export to Japan are put into water and the temperature of the water is decreased from an initial 24°C to 14°C (Samet et al., 1996). This is lowers their body temperature and incites fatigue. Once the prawns are in a calm state, they are packed in sawdust into 1kg boxes for air transportation to a Japanese market (which may take up to 36 hours). This practice works on the principle that the reduction in temperature slows the prawns metabolism, reduces its intake of O2, and therefore the likelihood of the prawn becoming stressed. An experiment by Samet et al. (1996) showed that the slower the reduction of temperature, the better the survival rate. Hewitt and Duncan (2001) found that in-transit mortality of P. japonicus occurred when prawns were harvested during periods of high temperature (>30°C ambient temperature).

5.5 Aquaculture practices

5.5.1 Managing aquaculture ponds

Pond management, together with pond water quality and sediment composition are the three most important factors in determining the success of aquaculture ventures (Funge-Smith and Briggs, 1998). Management of a pond is critical. Management activities involve: optimal stocking rates for size and depth of pond (as there is a correlation between high density stocking rates and smaller sized prawns); controlled feeding rates and quantities (to minimise organic waste); use and configuration of aerators (to ensure water in the pond

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management remains oxidised and homogeneous); water exchanges (to flush out excess nutrients and heavy metals); to encourage phytoplankton and bacterial proliferation (for toxic substance absorption); and, regulate ASS production and discharge (Wyban and Sweeney, 1989; Chien, 1992).

The following are a list of items that should be addressed when managing ASS in order to minimise the impact on the natural coastal and pond environments: reduction of ASS disturbance; leach acidity (by flushing pond with seawater to dilute and buffer acidic pH); neutralise acidity (with the use of limestone); in- situ burial beneath (clean) fill (to reduce the amount of DO diffusing into ASS); removal of anoxic storage of potential acid sulfate material (oxidise soil and treat leachate so it wont cause problems in the future); and, although not entirely economically viable, hydraulic separation of pyrite (Bowman, 1993; Sammut et al., 1996).

In some cases, farm management practices require the use of supplementary products to improve the water quality of ponds. They are generally in the form of:

5.5.1.1 Filters

Singh et al. (1999) undertook a study on four different filters to determine which was the most efficient in a recirculating aquaculture system. They tested the water properties of a bead filter, trickling filter, rotating screen filter and multi-tube settling basin. They found that the trickling biofilters preformed the best out of the four types for lowering the concentration of TAN (total ammonia nitrogen) and increasing the amount of DO in the culture tanks.

5.5.1.2 Managing Heavy Metals

One method which has been used to neutralise acidic pH, particularly in relation to discharge water, is the use of crushed or granulated limestone. This increases the pH of the water, making it acceptable (according to ANZECC Guidelines, 2000) for discharging into the estuary or coastal waterway. By directly liming the soil, the pH is increased, heavy metals precipitate and their detection in the associated water is minimised. However, Derome and Saarsalmi (1999) suggested that excess liming can promote the dissolution of heavy metals (such as Cu, Ni and Zn) from organic material which is Page 128

Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management immediately released into the water. This is caused by an excess in the release of organic complexing agents and an increase in the oxidation complexation potentials. With the application of the limestone, cations such as Cu, Ni and Zn were replaced by Ca, Mg and K. This enabled the heavy metals to become mobile in the soil pore water. They also suggested that the heavy metals associated with the organic matter accumulated because the heavy metals diminished the efficiency of microbiological activity. These authors concluded that repeated light liming with replacement macronutrients would assist in managing acidic soil containing heavy metals.

5.5.1.3 Erosion

Pond wall stabilisation techniques have been developed to remediate pond wall erosion, this is particularly important in ASS affected ponds. The growth of vegetation on pond walls stabilises the soil, encourages normal biochemical activity, and, (depending on where the plants are placed) act as a biofilter to remove nutrients from saline aquaculture wastewater.

In saltwater aquaculture salt-tolerant plants (halophytes) remove 98% of total nitrogen and 94% of inorganic nitrogen and 99% and 97% of the total and soluble reactive phosphorus respectively (Brown et al., 1999). In closed systems, seaweed is a good biofilter, however in open systems (such as tidal estuaries) the seaweed may become coated with fine sediment or not receive the required light.

In freshwater aquaculture, plants such as pineapples can be grown on the dyke walls. This provides a means of stabilisation and decreasing of erosion and liberation of metals from ASS. The sale of pineapples grown on the banks supplements the farmers’ income (Brinkman and Singh, 1982). Vegetation grown on the pond walls can reduce pond turbidity. On windy days sediment is blown in from the dyke walls and is held in suspension by the choppy pond water (Smith, 1996).

5.5.1.4 Synthetic Pond Liners

In some ponds, synthetic Ethylene Propylene Diene Monomer (EPDM) pond liners are used to combat the discharge of metals from sediments to the pond water. Horowitz et al. (2001) found that similar liners released toxic metals into Page 129

Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management the pond water. These metals were found to be toxic to both the aquaculture species (prawns) and nitrifying bacteria. In ASS conditions, the soil underneath the liner can become extremely toxic (Funge-Smith and Briggs 1998) and any mixing with groundwater can result in acid water (containing heavy metals) being discharged into local estuaries.

5.5.1.5 Chemical Use

Chemicals to improve the quality of water for aquaculture species are commonly used. Gräslund and Bengtsson (2001) looked at the chemicals and biological products used in the south-east Asian aquaculture industry. They concluded that farmers need to reduce the amount of chemicals that are used as they had created a risk to human health, the environment and to farm productivity. They also noted that some of the chemicals used, leave toxic residues that may remain in the ponds for years after their application.

Chemicals such as aluminium sulfate (Al2[SO4]3{14H2O}) or aluminium potassium sulfate (better known as alum) (AlK[SO4]3{14H2O}) are used to settle out the suspended colloids from the pond water. However, alum also has the dual effect of removing phosphorus from the pond water.

Liming ponds has the desired effect of increasing the alkalinity of ASS affected acidic water. Agricultural limestone (calcite, dolomite, calcium oxide or calcium hydroxide) is the most common form of liming agent. Liming is also suggested by Gräslund and Bengtsson (2001) to increase microbial activity in the sediment.

In south-east Asia, zeolites (aluminosilicate clay minerals which can adsorb or desorb molecules or exchange cations) are used to remove H2S and CO2 by adsorption and ammonia by ion-exchange (Gräslund and Bengtsson, 2001; Chiayvareesajja and Boyd, 1993).

Fertilisers are often used to promote the growth of algae in ponds and are generally in the form of chicken, pig or cow manure.

Pesticides are used to kill unwanted organisms such as fish, crustaceans, snails, fungi, seaweed, and algae (Gräslund and Bengtsson, 2001).

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Disinfectants such as calcium hypochlorite (Ca[OCl]2) and sodium hypochlorite (NaOCl) are used to remove pathogens from both water and sediment before and during the grow-out cycle. Formalin (or formaldehyde solution) is used as an antifungal agent in hatcheries and ponds and sometimes added to water to remove ammonia. Both hypochlorite and formalin are extremely toxic to aquatic organisms in incorrect doses.

Gräslund and Bengtsson (2001) suggest the following measures to be implemented to reduce chemical usage in aquaculture ponds:

• selection of locations for aquaculture where ponds are not impacted by acid soils which have low pH’s, and increase the potential for disease outbreak;

• proper design and construction of ponds to ensure they are neither too deep (could cause oxygenation issues) or too shallow (solar radiation could harm prawns and shallower ponds may mean warmer water and promotion of algal blooms);

• Implementation of sedimentation ponds to reduce the amount of suspended sediment and colloid that are released into estuaries. By treating the water outside the pond, the prawns wont be effected by alum and other chemicals used for sediment flocculation;

• Native aquaculture species selection to ensure the correct conditions for grow-out and reduce disease (as they should already have immunity);

• Reduce the amount of prawns in the pond and therefore reduce competition (and related stress) and increase the concentration of DO per prawn;

• Quality nutrition – improves growth and increases survival rates;

• Regular water exchanges and measurement of DO could assist in improving water quality and general health of prawns;

• Discharged water treatment before it enters estuaries and coastal areas to ensure pathogens, heavy metals and suspended solids are not released.

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5.5.2 Managing the Discharge Environment

Aquaculture, especially intensive farming, can degrade the adjacent terrestrial and aquatic environments. Some of the effects are: eutrophication and turbidity of coastal waters; mangrove destruction during construction of the ponds and drains; changes to the natural tidal patterns; disease from the farm spreading to estuarine organisms; exposure of ASS; and, salinisation due to land clearing (Wolanski et al., 2000; Burford et al., 2003).

There are two common ways of minimising the impact of discharged pond effluent to the environment: (1) feeding less or reducing the amount of water changes (allowing the nutrients to accumulate in the pond rather than the receiving waters); or (2) by improving the effluent before discharging (through the use of settlement ponds and constructed wetlands), or (3) both (Teichert- Coddington et al., 1999).

5.5.2.1 Natural Treatment

Effluent water can be treated by filtration. This is often with the use of natural or artificial filter. Oysters act as natural biofilters and filter out small inorganic particles, bacteria, phytoplankton, total nitrogen and total phosphorus from pond effluent (Gautier et al., 2001; Jones et al., 2001). However, high sediment loads can reduce or stop the filtration process of oysters. In a 1ha pond which undergoes a 20% water exchange each day and has a medium stocking density of P. japonicus, about 120,000 oysters are needed (and 12% of the total area of the pond should be allocated to the oysters – as each oyster is calculated to need 0.01m-2) (Jones and Preston, 1999).

+ However natural biofilters such as oysters also excrete excess NH4 , amino 3- acids, urea, uric acid and PO4 into the effluent water, but were able to remove excess ammonia from the water (Jones et al., 2001).

The use of oysters should be combined with regular water exchanges; seaweed (to trap suspended solids before released into estuaries); microalgal cultures (to absorb excess nutrients); and, bioaugmentation (the release of bacteria such as Nitrosomonas sp. and Nitrobacter sp. to assist in the

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management assimilation of nitrogen species); the use of settlement ponds and wetlands as biofilters (the latter which tackle both suspended solids and nutrient issues).

5.5.2.2 Settlement Ponds

Settlement ponds are used for both incoming and outgoing water at aquaculture farms. The incoming water (from the estuary) usually sits in a settlement pond to reduce the amount of suspended material that enters the pond and can be treated at this time if needed. This allows clean water to be pumped from the holding pond at any time (not reliant on the tides). Linking this process with biological filtration would ensure that the outflow water contains minimal waste.

Settlement ponds and wetland areas are also used to treat pond water before it is released into the estuary. For settlement to be effective the water needs to remain in the settlement ponds long enough for flocculation of suspended solids. Settlement ponds and artificial wetland preparation may be impracticable due to land size restrictions. The residence time for salt water is less than fresh water because the sedimentation rate of suspended solids is faster in saline water. If settlement ponds are aerated, the combination of high oxygen and ammonia concentrations attributes to the rapid nitrification of ammonia by aerobic bacteria. Underneath the sediment surface where oxygen diffusion is slow, conditions would rapidly become anaerobic (Chien, 1992; Briggs and Funge-Smith, 1994; Teichert-Coddington et al., 1999). If settlement ponds contain microalgae, then about 60% of dissolved inorganic nitrogen and dissolved reactive phosphorus is removed by absorption and mechanical means (Pagand et al., 2000).

Gautier et al. (2001) suggested that the following management considerations should be implemented to improve water effluent being discharged from an aquaculture farm: site selection; design and farm construction; management practices; and, effluent treatment.

5.6 Impact of Weather on the Pond Environment

There are marked differences on pond water chemistry between the summer and winter months. The main influences are: the length of the day (i.e. amount

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Chapter 5: Review: Aquaculture - Species, Water Quality And Pond Management of solar radiation); temperature; precipitation; wind speed; storm frequency; and increasing the amount of feed in the summer months.

Weather conditions are attributed to fluctuations in pond: salinity; temperature;

DO; dissolved reactive phosphorous (DRP); and, NO3-N. The last two elements were noted to change pond water quality by regulating the growth of phytoplankton (Cowan et al., 1999).

5.7 Summary

Successful aquaculture is dependent on selecting a farm location with soils that complement the farming technique and access to a sufficient quantity and quality of water. The species being farmed must be appropriate for the climate and water quality. Ideally, these species should be obtained locally to maximise the amount of natural resistance to disease. The farm layout should be prepared with water, waste treatment and handling in mind so as to preserve the adjacent environment as this is the farms lifeline to productivity.

Only once these planning components have been adequately thought through and accommodated can the farm manager be allowed to focus on the core business of maximum yields, minimising unnecessary cost and maximising profit for the good of all stakeholders.

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6 FARM MANAGEMENT/GENERAL NOTES FROM THE FIELD

6.1 Introduction

This chapter provides a short history of the development of the farm and gives a detailed account of the farm management procedures during the time of this study. Information presented in this chapter provides a background for the following hydrochemistry chapters.

6.2 Tomei Pond Logistics

The sequence of air photos presented below (Plate 18 - Plate 22) demonstrates the changing land use over the Tomei prawn farm and adjacent areas.

In the two earlier air photos there is evidence of cane farming and associated activity but little development on the Tomei prawn farm site (Plate 18 and Plate 19).

Simmonds and Bristow (1995) report that from 1980-1987 the western section of the Tomei prawn farm area was used for cane farming, and the eastern section was undisturbed salt marsh. In 1987, excavation of the eastern portion commenced from the south to the north as prawn aquaculture ponds were prepared.

The air photographs show that the farm was built in stages (Plate 20 and Plate 21). Plate 20 shows the initial excavation of the ponds in the southern end of the farm. By 1995 (Plate 21) the southern ponds were completely constructed and functional (containing aerators). This excavation led to the removal of the mangrove dominated setting. The ponds are still flanked to the east and north by Avicennia marina australascia mangroves. On the southern and western sides are sugar cane farms (Plate 22).

During pond excavation the dyke walls were built up using the excavated material from the ponds which was supplemented with road base (from the Stone Masters Quarries at Yatala) (Simmonds and Bristow, 1995). There is minimal vegetation (salt marsh grass and succulents) on the dyke walls (planted by the farm to decrease the effects of erosion: see Plate 23). Due to

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Chapter 6: Farm Management/General Notes from the Field the acidity of the soil, this vegetation has not thrived and the area remains very sparsely vegetated.

There are currently 20 ponds at Tomei (Plate 22), most of which are about 1ha in area and about 1.5m deep. Two ponds selected for this study are Pond 7 and 10 which are approximately 7,250 m2 and 14,200 m2 respectively (Table 9).

Plate 18: Air photo from 1944 which predates the establishment of the Tomei prawn farm, the 2002 farm outline is shown in yellow. Note that in 1944 most of the area was vegetated by mangroves.

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Plate 19: Air photo from 1970 which predates the establishment of the Tomei prawn farm; in 1970, there is little change in the land use pattern.

Plate 20: Air photo from 1990; the south eastern section was excavated in 1987 for the commencement of marine aquaculture activity, and the western section was being used for cane farming.

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Plate 21: Air photo from 1995; the southern ponds were being actively used for marine aquaculture, and the western section was still being used for cane farming.

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Plate 22: Air photo from 2002; the Tomei prawn farm was fully established and operational. This photograph was taken in February so coincides with the 2nd sampling round – note the distribution of paddle wheel and aspirators.

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Plate 23: (a) Pond 10 showing the irregular distribution of vegetation (salt marsh grass and succulents) on the dyke wall, (b) close up of more densely vegetated succulent growth on one of the pond dyke walls Table 9: Ponds 7 and 10 Dimensions

Pond # Pond Edge Length (m) North West East South Pond 7 105 87 97 61 Pond 10 81 167 113 140

Water level in the ponds varied regularly but the average pond depth in the centre of the pond was 1.9m in Pond 7 and 1.6m in Pond 10.

Low-grade, black, woven, synthetic material was used to cover the pond dyke walls to decrease the potential of erosion (Plate 24). The pond floors were left unlined, and a 100-200 mm layer of coarse-grained clean sand (Plate 25) was

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Chapter 6: Farm Management/General Notes from the Field spread on the bottom of the pond within which the P. japonicus were expected to bury.

Plate 24: Synthetic pond wall liner is used only on the dyke walls at the Tomei prawn farm

Plate 25: Imported sand is spread on the pond base (approximately 10-20cm thickness) prior to filling in preparation for the new growout season at the Tomei prawn farm

All water and soil samples were collected at the farm during the October 2001 – July 2002 season and due to their farming practices; the top soil in the ponds has been exposed to oxygen during pond excavation.

When installing the permanent mooring poles in the centre of Pond 7 prior to pond filling, the augured hole kept filling with water. It has also been noted by Eduardo Viso (pers. com. 2002) that water accumulates when king tides occur in the bay in the bottom of Pond 7 (when it is empty) and needs to be pumped out.

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At Tomei an algal bloom is promoted in each pond at the beginning of the grow-out cycle and that is maintained until harvest. The ponds are only about 1.5m deep, and to reduce the light intensity at the bottom of the ponds algal blooms provide shade and minimise prawn stress. The algal bloom is also a supplementary food source to the prawns.

The environmental manager, Eduardo Viso, indicated that ponds were stocked with Penaeus japonicus post larvae (PL) produced in the onsite hatchery. The P. japonicus PL seed was collected from a single supplier in Mackay. The farm also produced P. monodon and this seed is collected from Innisfail and Cooktown. In the 2001/2002 season, Pond 7 received 32.8 prawns/m2 and Pond 10 33.1 prawns/m2. The ponds were stocked in the first and second weeks of October 2001. The stocking densities in the 2001/2002 season were 237,756 (Pond 7) and 470,000 (Pond 10) prawns respectively.

After 8 months, at the end of the grow-out season, the ponds were harvested by the use of baited traps (collecting up to 3kg/night/trap in a good harvest) and then were drain-harvested to ensure all prawns were collected from the ponds (Viso, pers. com. 2002).

In the 1999/2000 season, the Queensland prawn industry produced 2301 tonnes of prawns (204.6 tonnes being Kuruma prawns) (Lobegeiger et al., 2001). At Tomei Australia Pty Ltd, total production for the same season was 52.8 tonnes as shown in Table 10, during the 2000-2001 season Tomei produced 18.9 tonnes and in the 2001/2002 season they produced 37.3 tonnes of prawns. In the latter season all 20 ponds were farmed and Pond 7 produced only 53 kg of prawns (0.2% of total production) and pond 10 produced 1550 kg (4.2% of total production).

The food conversion ratio (FCR) for the two ponds in the 2001/2002 season was: Pond 7 FCR = 137.0 and for Pond 10, FCR = 7.4 (Viso, pers. com. 2002). The FCR is calculated using Equation 10:

Equation 10 FCR = Dry weight of food fed (g) / Wet weight gain (g) (Sedgwick, 1979)

Where: FCR (food conversion ratio) = total weight of pellets eaten by prawns (g dry weight) / (divided by) Total prawn biomass increase (g wet weight) (Allan

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Chapter 6: Farm Management/General Notes from the Field et al., 1990; Allan and Maguire 1991; Briggs and Funge-Smith 1994). The lower the calculated FCR, the more productive the pond.

The following table contains summaries for both Ponds 7 and 10 for the 2001/2002 grow-out season and shows the differences between the ponds:

Table 10: Summary Table for Tomei Prawn Farm 2001/2002 growing season

Characteristic Tomei Australia Soil Type Clay (ASS) + layer of clean sand on pond bottom Distance from estuary (m) 20 Previous land use Mangrove/cane farming Source of water From adjacent estuary Number of ponds in farm 20 Type of cultivation Intensive prawn farm Aeration: Time in use (h/day) C 10 Type Paddle wheel and air diffusers Production duration (weeks) C 32 Pond 7 Size (m2) 7,250 Depth (m) 1.9 Number of paddlewheels 8 Number of aspirators 1 Stocking (# of post larvae) 237,756 Stocking Density (prawns/m2) 32.8 Feed (kg) 7,261 FCR 137.0 Total Production (kg) 53 Water exchange (m3) 162,862 Pond 10 Size (m2) 14,200 Depth (m) 1.6 Number of paddlewheels 10 Number of aspirators 8 Stocking (# of post larvae) 470,000 Stocking Density (prawns/m2) 33.1 Feed (kg) 11,509 FCR 7.4 Total Production (kg) 1,550 Water exchange (m3) 281,827

6.3 Water Exchange/Water Level Fluctuations

The water level in the ponds fluctuated each time a water exchange was undertaken. During high tide, water was pumped from the estuary into the farm settling ponds (Plate 22). Each pond received a 10%-15% (by volume) water

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Chapter 6: Farm Management/General Notes from the Field exchange about 4 times a week. Figure 19 shows that there were regular water exchanges in the summer months, this was done to maintain water level when there was greater evaporation from the ponds.

Figure 19: Volume and frequency of water in Ponds 7 and 10

The ponds have 8-10 boards below the water level, located around the pond for water exchange. If a pond was only having a small water exchange, then only one of the intake boards was opened, if a large water change was desired, then 3 or more boards were opened allowing >10% per day (Viso, pers. com. 2002).

Data supplied by Tomei Pty Ltd (Figure 19) show that water was exchanged in both ponds most frequently from December 2001 to April 2002. The greatest volume of water was exchanged in December and March for Pond 7 and December and January for Pond 10. Pond 10 water levels are more constant than the Pond 7 levels: this may be due to the difference in pond volumes or be a sign of farm management’s attempts to flush visible problems in Pond 7.

Water was pumped into and out of the farm from the same pipe and drainage channel (Plate 1). It is likely that there is an accumulation and recirculation of chemical constituents. Runoff from adjacent cane farms drain into the intake water channel and is likely to be modifying the chemistry of Tomei’s intake water through the potential introduction of acid (i.e. runoff), fertilisers, pesticides, heavy metals and agricultural chemicals used at the cane farm.

Water level dynamics were observed before the sampling was carried out and factored into the sampling design. To ensure consistency in the data set it was

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Chapter 6: Farm Management/General Notes from the Field important that a representative sample was taken from the same 3D point in the water column throughout the grow-out season. To allow for this a bolt was placed at the bottom of the permanent mooring pole (that held the multilevel piesometer in position). Each time samples were taken from the multilevel piesometer, they were taken from the same point relative to the pond base.

6.4 Aeration

During the 2001/2002 season ponds at Tomei are aerated with a combination of aspirator pumps and paddle wheel aerators (Figure 20). There were 8 paddle wheels and 1 aerator in Pond 7, while Pond 10 had 10 paddle wheels and 8 aerators. On average, water was exchanged four times a week and ranged from small/medium/large water exchanges depending on the health of the algal bloom in the pond. Prawn harvesting commenced in the first week of March 2002. It was not until the last week of June 2002 that the aerators were taken out of Pond 7 and the second week of July 2002 for Pond 10.

The effectiveness of aerators mixing the pond water at Tomei was tested by the author. A flow meter was lowered down to the sediment-water interface while paddle wheels and aerators were in use. The meter recorded little to no flow in this region of the ponds tested.

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Figure 20: Location of aerators, paddlewheels, water samplers and jetties in (a) Pond 7 and, (b) Pond 10

6.5 Data Collected by Farm Management

The following list of water quality variables were measured for all ponds at the farm on a twice daily basis by the staff at Tomei: water temperature, salinity (in the form of electrical conductivity or EC), pH, dissolved oxygen and secchi depth. These variables were measured using a handheld water quality meter (TPS) and a secchi disc. The depth of measurements were inconsistent (due to the variability of the people taking the measurement over the grow-out season) however, most of the data was reported to be collected from the top (approximately 500 mm) of the water column in each pond and always from the edge. This data may not be of a high quality (as there is some uncertainty about the frequency of calibration of the meters and the possible human error

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Chapter 6: Farm Management/General Notes from the Field that may be involved in the method of data collection) however, it is useful as a loose tool to gain insight on the temporal variation in Ponds 7 and 10.

6.5.1 Temperature

The water temperature data collected by Tomei for both Pond 7 and 10 are shown in Figure 21. The temperature data in Pond 7 are fairly similar to that in Pond 10 at the same period of time. In both ponds, the water temperature increases in December 2001 and gradually decreases in the cooler months (March, April, May). As expected, the afternoon temperature is higher than the morning temperature. The temperature in the pond water may have a bearing on the chemical reaction rates in the pond if they are temperature dependent.

Figure 21: Daily morning and afternoon water temperature data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10.

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6.5.2 Salinity

The salinity of both Ponds 7 and 10 are similar (Figure 22) but Pond 7 has a slightly lower salinity than Pond 10 (particularly after 13/2/02). This is probably management induced because in March 2002 there were a larger amount of water exchanges in Pond 7 than in Pond 10 (Figure 19). Pond salinity is lower in November/December and increases in January/February. This effect on salinity corresponds with the rainfall data presented in Chapter 2 (Figure 6). The larger amounts of precipitation in November/December reduce the salinity in the ponds during this time. Salinity increases in the ponds in January/February because the rainfall is comparatively less. Figure 22 shows that the ponds (particularly in Pond 7) have a higher salinity in the afternoon than in the morning. This may be due to increased evaporation of pond water throughout the day and is more pronounced on windy, hot, dry days (Figure 7 in Chapter 2 showed that the wind speed was greater in the afternoon).

Figure 22: Daily morning and afternoon salinity data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10.

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6.5.3 pH

There is a notable difference between the morning and afternoon pH data (Figure 23). In both ponds, the morning pH values are consistently lower (more acidic) than the afternoon values. In Pond 7, at the beginning of the grow-out season there was less of a separation between morning and afternoon measurements, however over time, the difference builds. Figure 23 is interesting as it shows that there are rapid pH fluctuations near the beginning of the grow-out season. Between 27/11/01 and 5/12/01, the pH increases from 8.06 to 9.68 (Figure 23).

Marine aquaculture farmers that work in ASS affected areas need to lime ponds (usually with CaCO3, but CaMgCO3 is also used, generally at the beginning of the season). This is to combat the effects of acid leaching from the sediment, maintain water alkalinity and in turn, to promote algal proliferation. Seawater does not have sufficient pH buffering capacity to combat the initial slug of acid that comes from the pond at the beginning of the grow-out seasons. After the addition of lime to Pond 7, a pH of about 9.4 was maintained for about 12 days. After that time, however, the pH decreased relatively rapidly to 7.94 (on 31/12/01). A similar trend appears in Pond 10 on a number of occasions.

Adding lime to ponds increases the alkalinity, however it may be more effective to add smaller amounts at regular intervals over the grow-out season to stabilise the ponds pH and minimise rapid pH fluctuations. Alternatively the farmer could add the lime to the pond bottom soil (mixed into the substrate) before the pond is filled with water (generally a better approach as often lime that is added to ponds remains in the water column and is lost during water exchanges). The lime assists with neutralising the ASS (before the prawns are added to the pond) and reduces the potential for acid discharge from the sediment-water interface while the prawns are still young and prone to suffer stress.

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Figure 23: Daily morning and afternoon pH data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10.

6.5.4 Dissolved Oxygen (DO)

Dissolved oxygen concentrations are consistently higher in the afternoon, and this is probably due to photosynthesising algae in the water column contributing O2 during the day. At night, when there is no sunlight to drive photosynthesis DO is consumed by algae and the organisms in the pond as they respire. This activity may also effect the pH of the pond, during the day the algae consumes CO2 and therefore the pond should be slightly more alkaline; at night the algae consumes O2 and releases CO2 (making the water more acidic at night). These reactions follow the general photosynthesis formula presented in Equation 11. The photosynthetic effects on DO can be seen in Figure 24, and are reflected by the morning pH readings (Figure 23) which are consistently more acidic than the afternoon measurements.

Equation 11 General Photosynthesis Formula 6H2O + 6CO2  C6H12O6 + 6O2

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Both ponds also show that the concentrations of DO at the beginning of the grow-out season were similar in the morning and the afternoon, however as the pond “matured” or ripened, the DO concentrations of the morning and afternoon plot further apart. This lag time was possibly due to the time it took for an algal bloom to proliferate and to effect the DO cycle in the ponds.

Figure 24: Daily morning and afternoon DO data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10.

6.5.5 Secchi depth

Secchi depth is a measurement of phytoplankton proliferation which is important as it provides sunlight protection to the prawns (Chien, 1992). Jamu et al. (1999) and Chien (1992) recommended that in summer a good reading is 300-400 mm and 200-300 mm in winter. At Tomei, the secchi depths average 500 mm in Pond 7 and 400 mm in Pond 10, but it peaks at 1000 mm in Pond 100 and 800 mm in Pond 7 (Figure 25).

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Over the duration of the 2001/2002 grow-out season, the ponds at Tomei underwent cycles of algal blooms and crashes. The cycles are related to the pond water chemistry and weather patterns.

Figure 25: Daily Secchi depth data provided by Tomei Farm Management, (a) Pond 7 and, (b) Pond 10.

6.6 Feeding regime

Feeding commenced in Pond 7 on Oct 12th 2001 and on the Oct 19th 2001 for Pond 10. The juvenile prawns were fed a high protein diet, four times a day at the beginning of the cycle. Initially the juveniles were fed a mixture of algae and artificial feed. When the prawns reached about 6g, they were fed once a day with approximately 8-10, 10kg bags of pellets. The pellets were scattered into the ponds from a boat to ensure that these were evenly distributed throughout the pond. The pellets used at the farm were a combination of Lucky Star and Higashimaru. The feed rates were calculated using feed charts

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Chapter 6: Farm Management/General Notes from the Field provided by the feed company (Viso, pers. com. 2003). These rates are based on optimum conditions; and, would not have taken into account of mortality due to influences from ASS. It is likely that the farm was over feeding the prawns

6.7 Bio-accumulation

The final harvest was carried out by draining the ponds and revealed a thick and diverse benthic community on the floor. Samples of the flora and fauna were collected for identification after draining in the 1999-2000 season. Fauna collected from the bottom of Pond 7 contained mud crabs, pippies, ascidians (Pyurra sp.), sea slugs (Bursatella leachi), polychaete tubeworms (Serpula rubens) and hydrozoans (Plate 26). Algal samples were analysed for metal accumulation. Gosavi et al. (2004a) who identified four dominant benthic species: Chaetomorpha; Cladophora; Enteromorpha; and Ulva. The Cladophora; Enteromorpha species contained metals in the following relative concentrations Fe>Al>Zn>Cu>As>Pb>Cd. Chaetomorpha and Ulva species had similar trends, however Pb>As. During the grow-out season as the ponds age, accumulation of waste occurs at the bottom of the pond. The waste contains material such as decomposing uneaten food, prawn and marine fauna excrement and decomposing microalgae. All benthic organisms in the ponds respire and therefore are sources and sinks of DO in the ponds and have a profound affect on the chemical reactions and concentrations at the sediment-water interface.

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Plate 26: Benthic fauna and flora, (a) single Bursatella leachi; (b) proliferation of Bursatella leachi; (c) egg cluster of Bursatella leachi; (d) iron stained tube worm cast Serpula rubens; (e) common example of the benthic organisms found on the pond floor at the end of the season.

6.8 Pond preparation

Each year after harvest, organic matter is removed by earthmoving equipment and the ponds are left dry until the following grow-out season (approximately 3

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Chapter 6: Farm Management/General Notes from the Field months). The pond base and walls are then disinfected (Viso, pers. com. 2001) in order to kill bacteria and organisms that introduce disease and impact the next grow-out season. Farm management indicated that the bottom of the ponds are sterilised with chloroform and then limed (to neutralise ASS). Once the ponds start being filled with water, calcium nitrate and di-ammonium phosphate are added to promote the growth of algae (Viso, pers. com. 2004).

The ponds were lined with a layer (up to 300 mm deep) of clean, coarse grained sand (Plate 27). This was intended for the burrowing Penaeus japonicus prawns to habituate.

Plate 27: Pond Preparation at Tomei

6.9 Disposal of solid waste

The organic matter and sediment that is mechanically removed from the bottom and walls of the pond is stockpiled. This material contains nutrients and oxidised and semi-oxidised sulfidic sediments and is a potential source of acid leachate. Before this material is relocated into the soil pit (located on the eastern rim of the farm near the creek), it is limed to raise the pH to neutral conditions. At the beginning of the 2000/2001 season, pond bottom sediments were dredged, the first time since the farm was established (Viso, pers. com. 2003).

6.10 Disposal of waste water

Water from the ponds is discharged into Jumpinpin estuary system and eventually into Moreton Bay (Figure 2). About 10% of the water is discharged

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Chapter 6: Farm Management/General Notes from the Field from the ponds each day during the normal water exchange process. Once the water leaves the pond, it flows through a series of drains until it reaches the discharge pipe (Plate 1). It is then discharged into a tidally controlled drainage channel and discharged into the estuary and then moves with the tide to Moreton Bay. The water discharges quickly and is naturally aerated upon discharge. This water is not treated upon discharge and contains relatively low pH and appreciable concentrations of trace metals (See Chapter 11).

6.11 Chemical manipulation of the ponds and associated uncertainty

During the grow-out season the ponds were treated with chemicals to promote the growth of algae and to crudely mitigate the affects of the ASS on water quality and protect against prawn disease. Eduardo Viso (pers. com. 2002) stated that agricultural lime and sometimes formalin is added to the ponds. Formalin is used for treating disease, and is only rarely used. On occasions, fertilisers such as urea, triple phosphate, and di-ammonium sulfate, is added to the ponds to promote algal growth. Burnt or slaked lime is added to the ponds if the pH drops very quickly.

There is therefore some uncertainty to which chemicals were used in the ponds and when they were applied (as the farm does not keep records of exactly what was put in and when – Viso, pers. com. 2002).

6.12 Summary

The ponds at Tomei were excavated in mangrove sediments between 1987 and 1995. Before the 2001-2002 growout season (the season that this thesis is based on) ponds were scraped, disinfected, limed (to help combat effects from ASS) and a 300 mm layer of clean sand was spread on the bottom of the ponds. P. japonicus prawns were grown in the two study ponds; Ponds 7 and 10. Farm management implemented techniques such as regular water exchanges, regular water quality measurements (such as pH, salinity, water temperature, secchi depth and dissolved oxygen concentrations) were carried out to assure good pond water quality and in turn, high pond yield. Unfortunately the yield from the ponds (particularly Pond 7) was relatively low

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Chapter 6: Farm Management/General Notes from the Field during this growout season. Data presented in Chapters 9, 10 and 11 provide an insight to why these two study ponds did not perform as expected.

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Chapter 7: Field And Laboratory Techniques

7 FIELD AND LABORATORY TECHNIQUES

7.1 Introduction

This chapter will discuss the field and laboratory methods used during collection and analysis of water and sediment data. A summary of the field trips undertaken, their purpose and samples collected is presented in Table 11.

Table 11: Summary of field data collected during study

Date Type No. Samples Collected 26th - 30th November 2001 Pond 7 and 10 water 68 2nd - 7th February 2002 Pond 7 and 10 water 72 3rd - 8th April 2002 Pond 7 and 10 water 67 19th - 21st June 2002 Pond 10 water 31 8th July 2002 Pond 7 sediment 11 cores 25th August 2002 Pond 10 sediment 11 cores 23rd Feb 2004 Seawater reference 3

7.2 Preparation for water sampling

After the end of each production cycle, the ponds are drained, faecal matter and food waste is removed from the sediment surface, and the ponds are dried and disinfected. At the beginning of the new season, the ponds are graded, sand is spread on the bottom Plate 27, and the aerator support poles are put in pre-determined configurations (designed by farm management to achieve maximum aeration and circulation).

In late September 2001, once the ponds reached the stage mentioned above, the author constructed the water samplers off site (Plate 28) and a field trip was undertaken to install the water column and pore water samplers at the farm (Plate 29). Three sampling designs were developed after a review of past and current sampling regimes. The fixed-point sampling devices provide a reliable method of collecting regular water samples from the pond water column, sand layer and underlying sediments.

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Plate 28: Construction of piesometers off site, (a) construction equipment; (b) Food grade plastic tubing (measured, cut and filters installed); (c) piesometer support poles in centre of picture with sediment sample elbows; (d) installed sampling point with filter sock installed; (e) completed piesometer (top in the background) with hammer as scale.

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Plate 29: Installation of water sampling devices

7.2.1 Design and Installation of Piesometers

As the focus of the study is on the water chemistry of aquaculture ponds associated with ASS, it was decided that it was important to collect not only water from the water column, but also pore water from both the sand layer and the underlying clay sediment. To be able to achieve this aim, three different types of water sampling devices were designed and installed.

7.2.1.1 Water Column Sampling Installation

To collect water samples from the water column, a multilevel piesometer design was adapted from the groundwater literature to enable sampling. Ten,

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Chapter 7: Field And Laboratory Techniques multilevel piesometers were constructed from PVC plumbers pipe, each was 2m in length and 65mm in diameter. Seven, food-grade, 5mm plastic tubes were taped to the outside of the pipe at nominated intervals as measured from the bottom of the piesometer (base of the pond) (Plate 30). These 5mm tubes were open at the intake point and had 300 mm of excess pipe at the top of the piesometer. They were numbered according to their depth above the piesometer base as per Table 12.

Table 12: Numbering configuration for piesometers

Pond 7 Sample ID Pond 10 Sample ID Depth (cm) 7/1-2 10/1/2 160 7/1-3 10/1-3 120 7/1-4 10/1-4 80 7/1-5 10/1-5 40 7/1-6 10/1-6 20 7/1-7 10/1-7 10 7/1-8 10/1-8 -10 7/1-9 10/1-9 -50 Note: The datum is referenced to the bolt at the base of the water column sampler (see Figure 26)

These tubes enabled water collection from the same point in the water column each sample round. The multilevel samplers were designed to be removable to inhibit biofouling on the filter points between sample rounds. A new multilevel piesometer was used for each pond, each sample round. This was to ensure minimal chance of contamination between sampling rounds.

Five PVC pressure pipes measuring 40 mm in diameter and 3 m in length were installed in both ponds to support the removable piesometers while sampling. This smaller diameter PVC pipe was capped at both ends then buried to a depth of 1 m with 2 m above the pond bottom surface. A stainless steel bolt was inserted at ground level to ensure that the removable multilevel piesometer was installed to the same datum during each sampling round (Figure 26).

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Plate 30: Plastic tubing taped to the outside of the conduit during piesometer construction

&  %    / #     '       *%& (   "*      '        + , )    -   .   $     '   

"# 

 $%&

Figure 26: Schematic of constructed water column piesometer showing bolt at base

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Chapter 7: Field And Laboratory Techniques

7.2.1.2 Sand Layer Pore-Water Instillations

Pore-water from the 300 mm thick sand layer on the bottom of the pond was sampled using permanent mini horizontal piesometers (Figure 27). The horizontal piesometers were constructed using 15 mm PVC pressure pipe, slotted, capped (at one end) and covered with a mesh filter sock (Plate 31). At the other end was an elbow and extension which was terminated by another cap that had a hole drilled in the top. Through this cap a piece of 5 mm (internal diameter) plastic tubing was inserted (which was left open-ended) inside the slotted conduit.

During construction, the 5 mm tubing was reinforced by being sheathed in 10mm tubing to ensure that throughout the season, the 5 mm sample tube was protected from punctures. The sheathed sample tubing was taped to the outside of the 3 m length of 40 mm pressure pipe. Three horizontal piesometers were installed at each sampling site, one as a sample point and the other two as back-ups in case of blockage. They were buried in the approximate centre of the sand layer (about 100 mm below the sand).

- ,+  &     %    / #      

*    

/  " %& '*

"       &      *%&      %&   

Figure 27: Schematic of sand and ASS pore water samplers

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Plate 31: Stages of construction of the sand pore water mini-samplers

7.2.1.3 Pond Base Pore-Water Installations

The 3 m length of 40 mm PVC pressure pipe was capped at both ends and slotted over a 400 mm interval from 300 – 700 mm below pond bottom surface for collection of pore water from the acid sulfate clay (Figure 27 and Plate 32).

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Chapter 7: Field And Laboratory Techniques

Plate 32: (a) cutting 40mm PVC pressure pipe into 3m lengths; (b) slotting the 40mm PVC intake section of the ASS clay pore water piesometer

7.2.1.4 Support Poles

Five untreated (3inch x 3 inch) hardwood poles were installed in each pond at the five sampling sites. These mooring poles aided in securing the boat while the sampling tubes were attached (Plate 33).

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Plate 33: Installation and use of wooden support poles during sampling in the ponds at Tomei

7.2.1.5 Sample Tube Extensions

A reel of nested sample tubes was prepared and tapped together; this was long enough to reach all piesometers from the pond wall where analytical equipment was set up. This tube was stored on a tube roll for easy handing (Plate 34).

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Plate 34: Tube roll containing plastic tubes which are attached to pond piesometer for sample collection on dyke wall

7.3 Water

7.3.1 Sample Collection

Water samples were collected from Ponds 7 and 10 on the 26th - 30th Nov 2001, the 2nd - 7th Feb 2002, the 3rd - 8th April 2002, and for Pond 10 only on the 19th - 21st June 2002. Water was collected with the aid of 10 vertical water column samplers, 10 horizontal mini-piesometers, and 10 slotted piesometers. Water was pumped from all tubes on the multi-level, mini and slotted piesometers with the aid of a vacuum pump and a capped conical flask (Plate 34).

On two occasions the tubing became blocked with algae, and it could not be unblocked, a water sample was not obtained for that particular sample point in that round. A total of 238 water samples (105 from Pond 7 and 133 from Pond 10) were collected.

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Three seawater samples were collected from outside the farm (Figure 2). These samples were from: 20m from the intake/discharge point in the channel outside the farm; 50m from the junction between where the channel discharges into the estuary; and, the junction between the estuary and the ocean at Jumpinpin Bar. The grab samples were collected on 23/2/2004 on a falling tide with the use of a dingy. These samples were treated and analysed as for the pond water samples. They were also used as reference seawater samples (in conjunction with literature samples) and formed part of the database for the “seawater line” on the geochemical graphs contained in Chapter 10.

7.3.2 Chemical Analysis

7.3.2.1 General Variables

Unstable physical variables such as pH, Eh (redox potential), DO (dissolved oxygen), temperature, and EC (electrical conductivity) were measured in the field using handheld, battery operated, Orion field meters.

Dissolved oxygen concentrations in the sampled pond water were measured using a battery operated, portable Orion meter (Model 810 A plus) in mg/L (ppm) by using an Orion field DO (type) probe was used. The DO meters were calibrated using a polographic DO solution.

To measure pH, an ORION model 290A meter, with a triode electrode. The pH electrodes were calibrated daily using standard buffer solutions from APS (Asian Pacific Speciality Chemicals Ltd) of 4.0 ± 0.02 @ 25ºC, 7.0 ± 0.02 @ 25ºC, and 10.0 ± 0.05 @ 25ºC.

Eh was also measured with an ORION model 290A meter, containing a platinum (Pt) redox electrode. A fresh measure of 4M KCl solution was placed in the electrode. The meter was calibrated against ZoBell’s solution (ZoBell, 3- 4- 1946) (0.003M K4Fe(CN)6 in 0.1M KCl and 0.02M K3Fe(CN)6 in 0.1M KCl) mixed in equal volumes to obtain a reading of +236 mV (+244.8 mV after compensating for temperature).

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7.3.2.2 Sample Collection, Storage and Analysis

Water samples were collected for cation, anion and trace element analysis. All samples were filtered using a 60ml syringe, a hand-held filtering unit and Millipore™ (0.45 micron) cellulose acetate filter paper.

7.3.2.2.1 Filtering

Sharma et al. (1999) suggests that Al and Fe particulate matter is small enough to pass through 0.45μm filter paper. Broshears et al. (1996) found that even if the chemical reactions had occurred at time of filtering, freshly precipitated particles (such as iron flocs (Fe(OH)3) less than 0.04μm in diameter) may still have been small enough to pass thorough a 0.10 μm filter. APHA standard methods (APHA, 1998) documented that 0.45μm is the accepted filter fraction for collection of water samples for hydrochemical investigation; and this methodology was adopted in the present study.

7.3.2.2.2 Cations

Cation samples were filtered into HNO3 soaked and Milli-Q (18) rinsed 60ml polypropylene bottles. The filtered samples were acidified using concentrated nitric acid (HNO3) to ensure chemical preservation. These were transported in eskies to minimise temperature fluctuation.

Upon returning from the field, the samples were diluted to 1:10 (due to their high salinity), and were then submitted for inductively coupled plasma – optical emission spectrometry (ICP-OES) analysis.

The ICP-OES analysis was undertaken by a Perkin Elmer Optima 3000 DV (following the USEPA Method 1620: Metals by Inductively Coupled Plasma Atomic Emission Spectrometry and Atomic Absorption Spectrometry – Dorothy Yu, pers. com. 2003). The machine was calibrated using multi-element standards. The calibration was accepted only when the linearity R2 was >0.999.

After the machine was calibrated, an Initial Calibration Verification (ICV) standard solution was run to check for accuracy and a blank was run 10 times to check precision.

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The detection limits of the method were calculated as 3 times the standard deviation of the results. Then a Continuing Calibration Verification (CCV) standard solution (a mid-point calibration standard) was run through the machine to verify linear calibration. The CCV was run after every 20 samples and at the end of the run to check for drift (Yu, pers. com. 2003).

The ICP-OES system presented the concentrations of various cations in solution in mg/L. The results were corrected for the dilution factor (1:10). Due to dilutions, some of the concentrations of elements were below the method detection limits (MDL) and therefore may be present in the sample, however are not in high enough concentrations to be detected by this method after dilution.

7.3.2.2.3 Anions

Anion samples were also filtered (using the 0.45μm Millipore™ filters) into

125ml clean polypropylene bottles (HNO3 soaked and Milli-Q (18) rinsed) at the point of collection. They were transported back to the laboratory also in eskies to ensure that they were kept cool prior to analysis once back at the university they were transferred into and stored in a refrigerator to reduce the detrimental effects of bacterial activity.

The concentration of elemental sulfur in the pond water was analysed by ICP- OES (the standard output of an ICP-OES). A concentration of sulfate was recalculated by multiplying the ICP-OES output by a conversion constant (S x

2.9959 = SO4 (mg/L)).

Titrations were used to determine the concentration of unstable anions in - 2- solution. Carbon dioxide (CO2), bicarbonate (HCO3 ), and carbonate (CO3 ) were titrated against standard solutions using colorimetric indicators. Total - alkalinity (as HCO3 ) was determined in the field by titration of 0.01M HCl against 25 ml of pipetted sample. Five to ten drops of bromcresol green or a methyl orange indicator were added to highlight the titration’s end-point (APHA, 1998). The end point occurred when all the bicarbonate converted to carbonic acid (H2CO3) at a pH of 4.3. This is chemically displayed in the reaction:

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- + Equation 12 HCO3 + H ⇔⇔⇔ H2CO3 ⇔⇔⇔ H2O + CO2(g)

Two titrations were carried out for each sample of pond water, and a mean value recorded. The concentration of bicarbonate in the sample was calculated using:

0&1 !  0& +  0&    + =2 340&1 5  5     Equation 13

A titration between 50 ml of sample and 0.02M sodium hydroxide (NaOH) was performed to analyse for CO2 (free). Phenolphthalein and metacresol purple indicators were used to determine the reactions end-point which occurred at a - pH 8.3 when CO2 is fully converted to HCO3 and a pink/violet colour change is seen (APHA, 1998). The reaction proceeds according to the following reaction:

+ - Equation 14 CO2 + NaOH ⇔⇔⇔ Na + HCO3

The concentration of carbon dioxide was determined using the equation:

&1 !  6 10  6 10   =2 34&1 '  '     Equation 13

Chloride analysis was undertaken in the laboratory upon return from the field. Chloride (Cl-) is a conservative element (APHA, 1998) and generally found in seawater and rainwater in the form of NaCl. It is one of the major inorganic anions (APHA, 1998).

Cl- was analysed in the laboratory using the argentometric method (pg 4-67 APHA, 1998) which is also referred to as a Mohr titration (Hem, 1989) (Plate 35). Before titration for chloride (Cl-) was carried out, the samples were diluted with Milli-Q water to a volume of 100ml (10ml of sample to 90ml of Milli-Q). This was because the water in the ponds is seawater and contains high concentrations of chloride. If the samples were not diluted, a large volume of concentrated AgNO3 would have been required to reach the titrations end- point. A 25ml aliquot of filtered, unacidified, diluted pond water was titrated against a standard solution of 0.141M silver nitrate (AgNO3) with 10 drops of yellow potassium chromate (K2CrO4) indicator to an end-point where a red

(Ag2CrO4) solution is formed. The resulting data were recalculated allowing for

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Chapter 7: Field And Laboratory Techniques sample dilution (value in mg/L * 10(ml) = actual value). Final value of Cl- in the sample is calculated by Equation 14:

Amount of AgNO (ml) x Molarity of AgNO x 100 Cl- (mg/l) = 3 3 Equation 14 25 x 10 (dilution factor)

Plate 35: Chloride colorimetric titrations carried out in the laboratory

7.3.2.3 Other Unstable Elements

Reduced ions such as ferrous iron (Fe2+), sulfide (S2-), and nutrients such as 3- 3- + nitrate (NO4 ), phosphate (PO4 ), ammonia (NH4 ), were measured in the field at time of collection, with a portable HACH DR/2000 spectrophotometer (Plate 36 and Table 13).

The HACH DR/2000 was used because “changes in components such as DO or CO2, pH, or temperature may produce secondary changes in certain inorganic constituents such as iron, manganese, alkalinity or hardness. Use time-composite samples only for determining components that can be demonstrated to remain unchanged under the conditions of sample collection, preservation and storage” APHA, 1998 pg 1-24.

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Table 13: HACH DR/2000 Standard Methods used for unstable elements.

Ion Method (HACH#) Sample Volume Wavelength (ml) (nm) Fe2+ Penanthroline (8146) 25 510 S2- Methylene Blue (8131) 25 665

NO3 -N Cadmium Reduction (8039) 25 500

NH3 -N Nessler (8038) 25 425 3- PO4 PhosVer3 (8048) 25 890

Plate 36: Reduced ions were analysed in the field with the use of (a) HACH spectrophotometer and, (b) colorimetric methods

7.3.2.3.1 Trace Elements (ICP-MS)

Trace elements (Ag, Al, As, B, Ba, Be, Bi, Cd, Ce, Co, Cr, Cs, Cu, Dy, Er, Eu, Ga, Gd, Ge, Ho, La, Li, Lu, Mn, Mo, Nb, Nd, Ni, P, Pb, Pr, Rb, Sb, Sc, Sm, Sn, Sr, Ta, Tb, Th, U, V, W, Y, Yb, Zn, Zr) were quantified with the use of a inductively coupled plasma-mass spectrometry (ICP-MS) unit. These samples were diluted 30 times for the analysis process. This was done in accordance with the advice of Beck et al. (2002): Saltwater samples must be diluted and total dissolved solids (TDS) should remain below 1000 mg/L when using ICP- MS. If this process is not undertaken, the large concentration of Na ions will damage the ICP-MS unit.

There are limitations however when using an ICP-MS to analyse seawater. Warnken et al. (2000) suggested that there is a reduction in the resolution of the ICP-MS, due to the presence of dissolved salts. Samples must be diluted; otherwise salt deposition on the instruments components (such as the torch injector tip) would decrease the machines accuracy. These authors found that during extended runs, salt blockage and increased plasma ion density for easily ionised elements such as Na were the contributing factors of inaccuracy.

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Pre-treatment by sample dilution prior to analysis, reduces the associated ion interference and minimises ion suspension. Dilution of samples reduces salt content to <0.2%, and minimises salt deposition on instrument components (Warnken et al., 2000).

Data resulting from the ICP-MS analyses was recalculated to allow for dilution, tabulated and plotted. Interpretation of plots of the hydrochemical data are used to assist in delineation of the dominant hydrochemical processes active at the two ponds studied (see Chapter 11 for results).

7.4 Sediment

7.4.1 Sampling sediment

Sediment sampling is typically undertaken by one of the following three methods: grab sampling, dredging and coring (Bufflap and Allen, 1995b). Coring is the best method of sampling if concentrating on the heterogeneous properties of a soil profile, as it causes the least amount of disturbance to the sample (Bufflap and Allen, 1995b). The best method for sampling an environment that is spatially variable, such as an estuary, is to take systematic samples along a transect (Jassby et al., 1997).

Based on this, the author has used a stainless steel sediment corer with an inner plastic sleeve and has sampled at and between piesometer locations (for further information on sediment sampling and figure on sample locations, see Chapter 8).

7.4.2 Sample Collection

Sediment samples (soil cores) were collected from Pond 7 one week after draining on 8th July 2002 and from Pond 10 three weeks after draining on 25th August 2002. Eleven samples were collected from Ponds 7 and 10, both at and between the fixed (mooring pole) water sampling sites (Figure 28).

The cores were collected with a plastic-sleeve, soil corer (approximately 1.25m in length) (from Dormer Engineering) resulting in a sample containing an undisturbed circular soil sample having approximately 40mm radius (Plate 37). Sample profiles were collected until the corer would not penetrate the pond

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Chapter 7: Field And Laboratory Techniques sediment. The soil corer was washed with pond water and then rinsed with distilled water (away from the next sampling site). Once the soil cores were removed (together with their plastic sleeve) from the corers stainless steel barrel, the ends of the plastic tube were tied off, ensuring no air was left in the core. The core was numbered and stored, and the core barrels plastic sleeve was replaced so that the next core could be taken. The soil cores were immediately frozen to prevent oxidation and transported to the laboratory for processing.

Figure 28: Schematic of pond showing the sediment sample collection points for Pond 7 and 10

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Plate 37: (a) Stainless steel sediment sampler and plastic sample sleeve; (b) one way inserts/valves and stainless steel sample cutter; (c) complete unit with pipe wrench

7.4.3 Soil Core Analysis

Prior to analysis the soil cores were removed from the freezer and defrosted on the laboratory bench (this process took approximately 3 hours).

A pH spear probe (TPS pH meter – WP80) was inserted into the core at 20 mm intervals, a measurement of the pH was recorded: providing a pH profile similar to those presented by Liaghati et al. (2003) (Plate 38).

The Tomei soil cores were sectioned (into 100 mm intervals). The sediment samples were oven dried at 80°C for 24 hours using the method of Morin and Morse (1999). Drying the ASS samples at 80°C reduces the risk of volatilising minerals such as greigite which may occur at 105°C (Sammut, pers. com. 2004). The weight of core before drying and the weight of the core after drying was recorded to determine the moisture content of the core segment. Visual descriptions and soil core data are located in Appendix 3.

Each sample spit was sectioned using the “quarter” system (Lim and Jackson (1982)). One quarter of the sample splits was sieved for particle size analysis (PSA). To determine the particle size of the dry sediment, it was sieved for 3 minutes on shaker at a frequency of 60rpm using a 63 micrometers sieve

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(BS410/1986 - stainless steel mesh and stainless steel frame) (Plate 39). Once the coarse and fine fractions were separated, each fraction was weighted using a Mettler PL 1200 balance (Plate 40) and the results are discussed in Section 8.5.3.

The second quarter was submitted to I. Wainwright (XRF Laboratory UNSW) for analysis of total organic matter using the Loss on Ignition (LOI) method at 1020°C and analysis with a LECO-CNS 2000 induction furnace analyser (with a furnace temperature of 1350-1450°C). This was used to measure relative percentages of total carbon, sulfur and nitrogen in the sediment (Morse and Cornwell, 1987 and Morin and Morse, 1999). Carbon and sulfur are detected by IR cells and nitrogen is detected using a thermal-conductivity cell. The instrument has variable detection limits, depending on the sample however they are about 10-20ppm for C and S and about 50-100ppm for N (Wainwright, pers. com. 2004). The machine is factory calibrated but is checked frequently with sulphamethazine.

The third quarter’s fine fraction (<63μm) was submitted for X-Ray Diffraction (XRD) to determine the mineral phases present in the soil samples. The analysis was carried out at the University of New South Wales in the School of Biological, Earth and Environmental Sciences. A Philips X`pert diffractometer with monochromatic copper K radiation (((Ave) = 1.5484 Å), 40 kV, 30 mA was used. The minerals were identified with reference to the International Centre for Diffraction Data Powder Diffraction File (release 2001) (Slansky, pers. com. 2004).

The final quarter was retained by the author as an archival sample.

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Plate 38: Spear probe inserted into the core for pH readings

Plate 39: Particle Size Analysis (PSA) equipment

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Plate 40: Weighing out sediment fractions after PSA

7.5 Hydrochemical Data

7.5.1 Computer Manipulation

Various software packages were used during this study and are displayed in Table 14.

Table 14: Software used during this study.

Name of Software Use Reference Microsoft ® Word ® Word processing Microsoft ® Word® 2002 Microsoft ® Excel ® Data compilation, Microsoft ® Excel® 2002 manipulation and graphs SPSS Statistical data SPSS 12.0.1 for Windows (2003) manipulation Copyright © SPSS Inc., 1989-2003 PhreeqC Interactive Thermodynamic (released: Feb 7th 2005) by D.L. version 2.11.0.148 modelling (1) Parkhurst and C.A.J Appelo. Surfer (Win 32) Modelling and isopach Version 6.04 (1997) maps Copyright © Golden Software, Inc., 1993-1997 Microsoft ® Drafting of figures Microsoft ® PowerPoint ® PowerPoint ® 2002 Copyright © Microsoft Corporation 1987- 2001 Endnote Version 6.0 Reference database © 1988-2002 Thomson ISI Research Soft

(1) Both raw data and calculated pE values were entered into PhreeqCI. The pE value was calculated using Equation 15:

Equation 15 pE = Eh/2.303RTF

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(where R= 0.001987 kcal/mol deg; T= absolute temperature; F= Faraday constant of 23.061 kcal/volt g eq) (Langmuir, 1997).

7.5.2 Quality Control / Quality Assurance

A number of methods were carried out before, during and after sampling in an attempt to minimise contamination and experimental error:

• Sample containers were washed with Milli-Q water (18) (Millipore Inc., Millipore Iberica), soaked in 10% nitric acid for 4 days and then washed again with Milli-Q (MQ) water before being placed upside down to dry.

• A new water column sampler was used for each sample run and each outlet tube was rinsed with pond water for 10 minutes before sampling. The tube roll was washed after every sampling period and 20% nitric acid was left in the tubes to inhibit bacterial activity between sampling runs).

• The field conical flask used for initial sample collection was acid washed at the end of each day. It was filled with the pond water at each tube interval and the waste was thrown away. For every sample, the second flask full was used for analysis.

• All samples for ICP were 0.45μm filtered to minimise the potential for clay particles to be trapped in the sample. Bufflap and Allen (1995a and 1995b); and Kotlash and Chessman (1998) suggest that filtration of sediment pore water samples is important to inhibit the possibility of fine sediment causing interferences during the analytical stage or, by absorption or release of trace metals occurring.

• A blank and a sample of the MQ ultra pure water was run through the ICP-OES and ICP-MS with the samples to ensure the quality of the results.

- - • All analysis carried out by titration (Cl , HCO3 , CO2) were performed twice. If the results were not in agreement, the titration would be done a third time as a check.

• Eh electrodes have variable reliability due to electrode poisoning: the field measured Eh values have been used as a guide tool to find a generalised trend.

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• A new plastic sleeve was used for each sediment core and contamination was kept to a minimum by ensuring it remained clean before sampling. The barrel of the sediment corer was taken apart and rinsed in the pond water and DI water before the next core was collected.

Two methods of validating the quality of the hydrochemical data were used: (1) charge balance calculation, and (2) TDS-EC ratio.

7.5.2.1 Charge Balance Error

A mathematical way of determining if the output from major ion analysis is accurate is by calculating the charge balance error (CBE). Charge Balance (Electro-Neutrality) calculations measure the analytical accuracy of each sample in the data set. It is based on the principle that any water sample must be in a state of electrolytic neutrality. That is, the sum of all cations must equal the sum of all anions. The CBE was calculated using Equation 16 (expressed in meq/L).

B Cations - B Anions CBE% (meq/l) = x100 Equation 16 B Cations + B Anions

Freeze and Cherry (1979) considered a charge balance error of ±5% acceptable and this has been adopted in this study.

Figure 29 shows the charge balance error for water samples analysed during the course of this study. Only one sample point exceeded ±5%. It was a sample from Pond 7, February with a CBE of -8.45%. This was checked and the error could not be identified so the value has been excluded from the interpretation. All of the Pond 10 data had a CBE of below ±3%.

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Figure 29: Charge balance error plotted against Total Dissolved Solids (TDS) for (a) Pond 7 and; (b) Pond 10.

7.5.2.2 EC / TDS

In general, the charge carrying capacity of any fluid will increase as the concentration of dissolved ions increases. This idea is represented by Equation 17:

Equation 17 EC x A = TDS Where the value of constant A varies depending on the water type from 0.55 and 0.8 (Hem, 1989)

This relationship is used by hydrogeochemists to determine the quality of groundwater chemical data, once the relation for a particular suite of samples (water type) is defined. This approach is often used by agriculturists that do not chemically analyse all samples, but rather use EC as a measure of water suitability for irrigation or stock use. The relation “A” for seawater in the Tomei ponds is 0.72 for both Pond 7 and 10 (Figure 30). This can be used by farm management to better manage pond salinity.

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Figure 30: Relationship between TDS and EC for (a) Pond 7 and; (b) Pond 10.

7.5.3 Statistics

Descriptive Statistics (mean, standard deviation, maximum and minimum concentrations etc) were calculated using SPSS v. 12.0.1 for Windows. Spearman’s rank correlations were used in preference to the simple linear Pearson’s correlations. Correlating the data assists in determining the relationships between different metals and assisting in the interpretation of which chemical processes are important in the ponds. As pointed out by Liaghati et al. (2003), metal oxides (iron oxyhydroxides and manganese oxides) and organic carbon commonly act as scavengers for heavy metals. Therefore, complex chemical reactions related to metal associations may be understood by looking at correlations between metal oxides and particular heavy metals. As in this study they used the Spearman’s rank correlations because the data was not normally distributed.

7.6 Summary

During this study a unique dataset has been collected. The author selected and developed a method for collecting a network of samples over a season that is correctly spaced to allow interpretation of pond water quality through space and time. Detailed field measurements of unstable variables and both field/laboratory analysis using the current and reliable techniques provide an interesting and unique dataset. This study is the first time aquaculture ponds built in ASS’s have been studied in such detail.

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8 POND SEDIMENTS

8.1 Introduction

Even though the emphasis of this thesis is to understand the hydrochemistry of these ponds, the author believes it prudent to further characterise the pond sediments of Tomei as understanding the composition and reactivity of these sediments is critical for meaningful hydrochemical interpretation. Section 8.5 discusses the results of the author’s soil core investigations. Gosavi (2004) studied the chemical composition of the pond sediments at Tomei. Gosavi (2004) concluded that the pond bottom sediments at the farm are accurately classified as being ASS’s. Additional characterisation of the soils, their composition, grain size, texture and descriptions are provided by: Isbell (1996); Rassam et al. (2002); and Preda (1999).

8.1.1 Relationship between pore water and the adjacent sediments

Sediment pore water chemistry provides insight into the chemical composition of the sediment and vice versa. Calmano et al. (1990) noted that due to the small interstitial area between the grains in fine grained sediments minor chemical reactions in the sediment (solid phase) often translate in major compositional changes in the pore water (aqueous phase). Chemical characterisation of pore water and adjacent sediment provides evidence to support geochemical interpretation which may indicate that sediment has been influenced by chemical reactions that may be driven either by natural and/or anthropogenic activity.

Avnimelech et al. (1984) found that the sediments underlying a water body, contain more organics and nutrients and hence more microbes (e.g. decomposers) than that found in the water column. Just as pond sediments absorb nutrients and trace elements they, through reversible reactions, are able to release them. That is, sediments can be a chemical source or sink depending on changes in conditions. Organic matter and heavy metals that are in the sediments are respectively broken down and released into the pore water during anaerobic degradation (Lapp and Balzer, 1993).

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8.1.2 Impact on aquaculture species

Chemical reactions and the associated ion cycling between pore water sediments and the water column can impact the health of the aquaculture species. Ponds built in ASS’s have a greater potential (than those built in more inert sediments) to be detrimental to the culture species. Common problems experienced by farmers who have ponds that are built in ASS are: high mortality, slow growth, low yield, fast mortality from Al or Mn toxicity, depletion of P (as it is easily exchangeable with As) and the accumulation of organic acids (Golez and Kyuma, 1997).

Boyd (1992) noted that “the occurrence of reduced soil or sediment in a pond bottom is normal and not harmful to shrimp as long as there is a well oxidised surface layer on the pond bottom” (pg. 168). This statement does not hold true in ASS environments. A well oxidised surface layer on the pond bottom will accelerate reaction rates in ASS leading to rapid changes in pond chemistry which in turn is harmful to the aquaculture species. The thesis demonstrates that the reactivity of the ASS (even at only 300 mm depth) affected pond water between the water column and the sediment (with ions moving across). There is a perpetual exchange at the sediment-water interface, as a conduit for transport of ions between water and sediment. In Pond 7 at Tomei there is hydraulic connection to the adjacent estuary: tidal movement results in a rhythmic up-welling and discharge of ASS leachate into the pond.

8.2 Classification of Soil Type at Pimpama

Soils comprise: primary and secondary minerals (from the weathering rock and/or transport of sediments) and textures (sand, silt, clay), organic matter (living and dead), and in pore space of both water and air (in pore spaces).

The soils at Pimpama were classified as being an Ultisol or Sulfuric Hydrosol by Isbell (1996): they have an average clay content of about 30% and a medium particle size (D50) of about 0.0055mm (Rassam et al., 2002).

8.3 Clay Types

The dominant clay types in the Pimpama area are kaolinite, illite and smectite (Preda, 1999). These clays are secondary weathering products of the

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Palaeozoic Beenleigh Block Metamorphics which include sandstone and shale. Preda and Cox (2002) suggested that alteration of volcanic material may be the source for the abundant mixed layers of clay (illite and smectite) in the Pimpama region.

8.3.1 Soil composition and clay mineralogy

Gosavi (2004) analysed pond sediment samples at Tomei for sediment mineralogy (for whole soil) and found that 18-24% (weight %) of those samples were quartz. Appreciable quantities of albite and anorthite were also noted in the sediment matrix. Furthermore, the dominant clay mineral was in order of greatest abundance: mixed layer illite-smectite; illite; and, kaolinite. However findings by Preda (1999) and the XRD data presented in this thesis suggests that the relative abundance of clay minerals is in reverse order to Gosavi’s (2004) findings.

The percentage of kaolinite, illite and mixed layer illite-smectite contained in Gosavi’s (2004) samples increased with depth in all but one sample. Jarosite which is commonly associated with ASS was present in most soil samples excluding the deepest parts of the soil profile. The concentration of jarosite was the highest in the top 1m of the soil profile. This mineralogy change is best explained by pyrite deeper in the profile not being exposed to oxygen and hence not oxidising to form jarosite: deeper in the profile the sediments are PASS’s not ASS’s.

8.3.2 Chemical composition, structure and attributes of key clay minerals

Kaolinite (Al4Si4O10(OH)8) is a hydrous aluminium silicate. It is typically very soft and is commonly found in crystals or small pseudo-hexagonal plates (Read, 1984). It is the alteration product of feldspars which are associated with granites (Read, 1984). As kaolinite has a 1:1 (Al:Si) structure, it does not posses an electric charge and therefore, it does not generally bond with other non-clay species (Preda, 1999). It can have impurities (Ti, K and Fe3+) and accordingly can posses a slight electrical charge. Kaolinite generally has a low CEC and is not a significant contributor in ion-exchange reactions.

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Illite (KAl4(Si,Al)8O20(OH)4), a hydrous potassium aluminium silicate, is the most common clay in nature, and belongs to the mica group of clays (Read, 1984; Drever, 1997). It possesses a sheet structure and is formed by either weathering or hydrothermal processes and is the product of muscovite or feldspar alteration (Read, 1984). Illite is formed in low temperature conditions, has a structure similar to muscovite, and may contain some Mg and Fe. The CEC of illite is moderate and lies in between the smectite and kaolinite clays.

Smectite (or “swelling clay”) usually occurs as small crystals and is difficult to detect using XRD techniques. One method of detection is by applying ethylene alcohol to a sample, which expands the clay structure. Smectite clay contains water in the interlayer spacing of its 2:1 structure. The clay swells when the interlayer cation, commonly smaller divalent ions (Ca2+ or Mg2+) becomes saturated with monovalent ions (e.g. Na+ or K+) (Drever, 1997). These mixed layer clay minerals also shrink when the amount of water in the interlayer decreases. This makes the clay unsuitable for use in aquaculture in pond dyke walls. Smectite may be Fe3+-rich (nontronites) or Al-rich (montmorillonite- beidellites). Smectite has a large cation-exchange capacity (CEC); the greatest compared to other clays such as illite and kaolinite. Therefore it is significant in ion-exchange reactions. However, the CEC decreases when pH is low (such as in the case of ASS’s) and increases where the concentration of organic matter is high.

8.4 Environmental Effects – adaptation and chemical interaction

Mangroves are commonly found in estuarine environments where sediments are water saturated clay rich and are strongly reduced. Reduced sediments in estuarine environments are a source of H2S, N2O and other biogenic gases, these gases are typically released directly into the atmosphere (Jorgensen, 1983).

Tam and Wong (1993 & 1995) found that most of the heavy metals added to mangrove soil from discharged sewage were bound within the sediments by: adsorption to ion-exchange sites: incorporation into lattice structure; or, by precipitation as insoluble sulfide minerals: and concluded that only a very small

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Chapter 8: Pond Sediments portion of the heavy metals were bio-available. Tam and Wong (1997) studied the effects of wastewater (sewerage) on mangrove systems. They found that the roots of the mangroves trees contained higher concentrations of metals than in the aerial parts of these plants, and suggest that the roots act like a barrier to the metals (protecting the more sensitive parts of the plant from metal contamination).These authors found that metal uptake by plants depends upon factors which include: pH and redox conditions; plant species; changes in the seasonal growth rates; and, the age of the plant.

8.4.1 Impact of pH

Although some species (most probably those that have adapted to reduced or acidic environments) are able to block the uptake of heavy metals, some marine flora and fauna have not developed this adaptation and instead readily uptake heavy metals. This uptake results in heavy metal bioaccumulation in the food chain (Liaghati et al., 2003).

In acidic environments, elements such as Mo, P, Zn, Co, Mg, and Ca are less biochemically available to plants. Elements such as Al, Fe, and Mn become more available and in high concentrations, can cause plant toxicity and death. When pH is greater than 7.5, Ca can bind with P, making it less available to plants. The bioavailability of elements can stunt plant growth and reduce the yield of crops and pastures.

Golez and Kyuma (1997) investigated the influence of pyrite oxidation and soil acidification on the availability of selected nutrients. They found that repeated wetting, drying and oxidation of soil, decreased the concentration of pyrite and 2- reduced the measured concentrations of Fe, Al, SO4 . Strongly acidified soils were shown to decrease the solubility of Ca, Mg, Zn and Cu. Manganese was found to dissolve from the soil matrix under intensely acidic conditions; the resultant leachate was found to have a toxic effect on flora and fauna. Continued oxidation of ASS decreased the availability of P making the soil nutrient deficient. These authors concluded that cyclical leaching and flushing with seawater is one of the most promising methods for remediating ASS.

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8.5 Depositional Environment

Preda (1999) suggested that the Pimpama River valley and floodplain was flooded by seawater during the Holocene transgression (see Section 2.3.1). This is when the authigenic sedimentary pyrite that results in the regions PASS is suggested to have formed. Preda (1999) states that pyrite precipitation occurred through the input of: sulfur from the seawater; the availability of organic matter; input of iron from basement rocks; and the established conditions suitable for bacteria which catalysed chemical reactions and resulted in the formation of iron sulfides. Preda (1999) further suggests that the sediments deposited during this period, were deposited in cycles and resulted in different mixes of clay minerals dominating the soil horizons.

During sedimentation of this sequence the first depositional episode occurred in a low energy, estuarine environment, with smectitic sediments being deposited in incised river valleys.

The overlying kaolinite-rich sediments were found to be deposited in a disordered manner. Preda (1999) suggested that this was due to intense reworking of the material. This may have occurred during a flash flood or storm event. During this event a sandy, kaolinite dominated sediment was deposited over the smectite sequence. Preda (1999) also suggested that the sands were transported northward by longshore drift and entered the bay through the Jumpinpin tidal inlet.

During the highstand that followed the transgression, sandy kaolinitic clays were deposited in the west, and sand-rich, slightly kaolinitic units were deposited in the east (Preda, 1999).

Preda (1999) concluded that the Holocene and younger sediments consist of reworked Palaeozoic material deposited in an estuarine environment. Preda (1999) also noted that there was a fluvial influence in the western part of the area.

X-ray diffraction by Preda (1999) and Preda and Cox (2000) and LECO (for total sulfur), ICP and XRD by Preda and Cox (1998b) showed that the dominant clays at Pimpama are kaolinite, irregular mixed-layers of smectite-

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Chapter 8: Pond Sediments illite with lesser amounts of montmorillonite and illite. They found that the chemical compositions (expressed as oxides) were: SiO2 46-57%, AlO 15- 25%, FeO 10-15%, NaO 1.5-5%, MgO and KO both 1.5-2%, and CaO and TiO both 1-1.5%. Total S varied from 2.2% at the surface to >10% in the lower sections. The sediments also contained quartz and feldspars (derived from the Palaeozoic bedrock) and pyrite which were typically associated with jarosite and gypsum. The clayey samples were found to contain more pyrite than the sandy samples, and kaolinite was dominant over smectite.

Preda (1999) suggested that differential tidal energy that triggered preferential deposition of clays may also favour a heterogeneous distribution of organic matter and therefore pyrite. Scanning Electron Microscope (SEM) performed by Preda and Cox (1998b) showed that pyrite was in two forms; round framboids (~5μm) and elongated clusters of authigenic crystals (~50μm).

8.5.1 Relationship between depositional environment and metal mobility

When pyrite undergoes oxidation, the resultant acidity breaks down the clay structure and releases weakly bonded metals. Metal mobility is greatly enhanced in an estuarine environment by wave and tidal activity. The mobilisation of heavy metals is of concern, as the concentration of metals is typically above concentrations recommended as safe upper limits in the ANZECC Guidelines (2000). The high mobility, transport and deposition of these metals into new environments can introduce toxicity into otherwise healthy environments. Preda’s (1999) analysis showed that all the trace metals that were found in the deposited sediments (V, Cr, Mo, Co, Ni, Cu, Zn, and Pb) were also present in the bedrock. Zinc, Co and Mo were found to be higher (had been concentrated) in the sediments. Preda (1999) suggested that the Mo may have accumulated in the deposited sediments due to its low mobility in acidic environments; the Co and Zn likely accumulated and were attributed to anthropogenic sources.

8.5.2 Site Description – Soil Characteristics

Gosavi (2004) was carried out a study on the pond base sediment in Ponds 7 and 10 during the 2000/2001 season. Gosavi (2004) found that the pond sediment was dominated by eroded pond wall material, and that the pyrite

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Chapter 8: Pond Sediments disseminated throughout the pond dyke walls generated and introduced acidity to the ponds. Oxidation of the pyritic material was shown to have occurred throughout the entire dyke mound and was not restricted to the surface layers of dyke walls (Gosavi, 2004).

At the start of the 2001/2002 season, about 500 mm of the pond floor sediments were mechanically removed. This was in addition to the annual practice of using an excavator to scrape away built up organic matter from the bottom of the pond. Following this excavation work both Ponds 7 and 10 (at the start of the 2001/2002 season) had a 0-0.3° slope, and are therefore is depressional to flat.

The farm is built on estuarine sediments that hosted a substantial mangrove system. Mangrove soils are clay-rich and have a high organic content (usually 2-3 times higher than non-mangrove clay sediments).

8.6 Sediment Cores

Sediment cores were collected by the author to a depth of 575 mm. The average maximum core depth in Pond 7 was 452 mm, and 234 mm in Pond 10. Sediment cores were collected using a plastic sleeve lined, manually operated, one-way valve corer after the ponds were drained at the end of the growout season. The cores were collected from 11 sample sites in each of the two ponds; however, data presented in the following sections are only from the cores taken near the pond water samplers. The remaining cores were sieved; the lab pH was measured; and the cores were oven dried for preservation. Samples remain in the possession of the author and are available for further analysis.

The sediment core depths were limited by the maximum depth the corer could be pushed into the sediment. The cores from Pond 7 are longer than those collected from Pond 10: this is due to the refusal depth being shallower in Pond 10 (likely to be due to higher compaction and lower moisture content).

Future detailed work on the sediment in the pond, should consider using a vibrocore to capture a deeper profile. Rayment and Higginson (1992) classify deep sediment sampling as that from about 600-1000 mm, and suggest that

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Chapter 8: Pond Sediments sampling from this zone assists in a better assessment of sediment salinity, acidity, S, and the mineral N content.

8.6.1 Sediment Core Description

In ponds 7 and 10, the top of the core profile consisted of coarse grained sand (sand layer installed by the farm managers as part of their pond preparation methods.

After collection, the sediment cores were frozen until analysis. Visual inspection of the cores showed that the cores were water saturated and contained black mud (monosulfides) with a strong smell of H2S (Plate 41). These cores were not oxidised and remained unripe as defined by Dent (1986) and are good representative in-situ samples. Their unoxidised state was most likely due to their collection directly after the ponds were drained.

Plate 41: Black monosulfide contained at the base of a sediment core from Pond 10 is typical of those collected from Tomei. The sand layer at the top of the core is added during pre-season preparation and shows evidence of iron oxide staining.

8.6.2 Results from XRD

The XRD analysis was carried out by Dr Ervin Slansky (BEES, UNSW) in April/May 2005. The sediment was processed before submission for XRD with the fine fraction (<63μm) separated from the coarse by sieving. This was done to reduce any interference from coarse grained sand in the samples. The samples were dominated by detrital rock forming minerals such as quartz and feldspars (mostly plagioclase, and sometimes accompanied by alkali feldspar).

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There were also trace amounts of pyrite, halite, jarosite and gypsum. Some other minerals such as carbonates, ferrihydrite, bassanite, alunite, maghemite, copiapite, sphalerite, sylvite, anhydrite, smectite and gibbsite were identified as contributors to the various peaks identified on the XRD traces. Quartz and feldspar peaks are dominant and masked definitive identification of these accessory minerals (Slansky, pers. com. 2005).

In summary, the dominant mineral in both ponds was quartz (about 40-60%). Pond 10 contained a slightly greater concentration of quartz than Pond 7. Feldspar (20-40%) was the next most abundant and minerals such as pyrite, halite, kaolinite found in small amounts (5-20%) and jarosite and gypsum in trace amounts (<5%) (Appendix 3) (XRD traces are available from Ervin Slansky at UNSW XRD laboratory).

Preda and Cox (2001) analysed the minerals in the bedrock at Pimpama and found that the dominant minerals were: quartz, feldspar (plagioclase and K- feldspar), micas, pyrite, illite and interbedded smectite-illite.

8.6.3 Sand and clay fraction analysis

Particle size analysis was carried out on sediment cores taken in close proximity of the sediment pore water samplers. The descriptive statistics are recorded in Table 15; Appendix 3 provides the tabulated raw data. The author recognises that anything measured as <63 microns is a mixture of clay and silt (mud and fine material). Anything that is referred to as “clay” in this thesis is actually a mixture of clay and silt.

Table 15: Descriptive statistics for the particle size analysis for samples from Pond 7 (n=24) and Pond 10 (n=14)

Pond 7 Pond 10 Pond 7 Pond 10 % Sand % Sand % Clay % Clay Max 97.73 98.84 5.88 11.66 Min 94.12 88.34 2.27 1.16 Average 96.42 96.45 3.58 3.55

These data demonstrate that both Ponds 7 and 10 are dominated by coarse grained material (88.34 – 98.84% of the sample) and also contain a small proportion (ranging from 1.16 – 11.66% of the sample) of fine particles (including silt or mud).

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Figure 31 shows the graphical representation of grain size distribution for the sediment cores from Pond 7 and Pond 10. This analysis was carried out to determine if there was notable difference between the two ponds in relation to the proportion of sand and clay. Both ponds are shown in Figure 31 to have equivalent proportions of sand and clay, accordingly there is another reason for the geochemical difference between the ponds.

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Figure 31: Sediment core data summary for Pond 7 and Pond 10 (sand/clay ratios, moisture content, loss on ignition, results from the LECO analysis, and laboratory based sediment pH data)

It is important to recognise that the high proportion of sand in both ponds is most likely attributable to the sand layer that was applied as part of the farm management technique at the start of the season. This sand has mixed through the cores at the time if the sampling, or during the season through bioturbation.

8.6.4 Water Content

The water content in the sediment cores was calculated using Equation 18:

Total wet weight of sediment - total dry weight of sediment Water content = x 100 Equation 18 Total wet weight of sediment

(after the core was put into the oven at 80°C for 24 hours). These data are presented in Appendix 3.

Figure 31 (water content in the sediment cores) shows that the water content correlates with a higher proportion of clay. In ASS’s the clay material is commonly water saturated, and pore water can only be extracted by centrifugation: gels are also commonly recognised in ASS environments and dry to residual clay sized particles.

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8.6.5 Organic matter in Sediment Cores

The organic content of the pond sediment was measured using Loss on Ignition (LOI) analysis. LOI is a measurement of the total mass loss after incineration at 1020 °C (described in Section 7.3.2).

On average Pond 7 lost 6% more organic matter on combustion than Pond 10 (with the exception Pond 10 core 8). Figure 31 shows the LOI data from Ponds 7 and 10. Mangrove sediments have a high organic content which could be 2-3 times greater than clay sediments. It may be the case that Pond 7 had a slightly higher percentage of organic matter (compared to Pond 10) because of its proximity to the mangroves.

8.6.6 Sediment pH

Hydrogen ion activity (pH) was measured at 100 mm intervals to generate a pH profile for each core. This was done in the laboratory by inserting a spear probe into the defrosted core (pH reading was taken when the core reached room temperature). A similar process was undertaken by Liaghati et al. (2003). The sediment pH from the two ponds are presented in Appendix 3 and in Figure 31.

Figure 31 shows that the two sediment cores collected from near the pond dyke walls have a lower (more acidic) pH than the cores collected from the central (deepest) part of the pond. Sediment cores 4, 6 and 8 from the pond display a zone from 120-420 mm where the pH drops to approximately 7. In cores 2A and 10 (the cores close to the dyke walls) pH falls to 4.1 and 5.3 respectively. This lower pH correlates with high sulfur as determined by the LECO analysis.

In Pond 10, the cores taken from near the dyke walls (core 4 and core 10) have a slightly higher pH (are more alkaline) than the cores taken from the centre of the pond (Figure 31).

The pond water pH data (spatial pH data for Ponds 7 and 10 presented in Section 9.1.1.1 and 9.1.1.2) shows a good correlation with the sediment pH data. Figure 31 suggests that pH away from the dyke walls is less affected by ASS than the samples located adjacent to the dyke wall. Gosavi (2004)

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Chapter 8: Pond Sediments measured the soil pH in the dyke walls of the ponds at Tomei and found that the dyke slope soil pHsol ranged from 3.12-3.92 at 0.5m and 2.70-3.80 at 0.3m.

Gosavi (2004) also found that the mean pHsol for each of the slope profiles taken from the dyke walls increased down the wall towards the pond bottom. Gosavi (2004) results reinforce that a large amount of the acidity in the ponds is attributed from the dyke wall material and that the sediment at the bottom of the pond is not as acidic as the sediment in the wall material.

Data from Pond 7 display that the pond bottom sediments are not as acidic as the sediments at the bottom of Pond 10. It is possible that there is a tidally influenced estuarine intrusion buffering the sediment in the centre of the pond (when drained, the centre of the pond was always under water in Pond 7).

Pond 10 has a more uniform sediment pH. Pond 10 was noted to drain completely when empty so is not believed to be affected by estuarine water influx and pH buffering (Section 9.1.1.1and 9.1.1.2).

8.6.7 Carbon, Nitrogen and Sulfur in Sediment Cores

The C/N/S concentrations were measured in the sediment core using a LECO analyser (see methods in Section 7.4.3).

Comparing Ponds 7 and 10, on average there is 38% less C in Pond 7 sediments than in Pond 10 (Figure 31). There is also approximately 50% less N in Pond 7 than in Pond 10 (excluding the top 0-100 mm of Pond 7 core 4). There is a clear difference in sulfur proportion between Ponds 7 and 10 with Pond 7 containing on average 15% more sulfur than Pond 10. Pond 7 contains a slightly larger concentration of pyrite than the sediments in Pond 10 and therefore this is the likely source of sulfur. It is likely that the nitrogen in Pond 7 is being taken up by the aquatic organisms and is being converted to organic acids, and does not remain in solution.

8.7 Key Points to retain when considering the water chemical data

The rock forming minerals undergo weathering and/or dissolution due to chemical, mechanical and biological reactions that are occurring in the sediment. These minerals deliver cations, anions and trace elements (including

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Chapter 8: Pond Sediments heavy metals) to the pore water and subsequently the water column. Trace elements are dispersed into the water column through the sediment-water interface.

The key messages to retain and integrate into the hydrochemical interpretation are:

• The grain size is equivalent in each pond and is mainly coarse sand.

• Organic matter content is equivalent in both ponds.

• There is a direct relationship between high sulfur content and low pH.

• The main minerals are quartz and feldspar with some pyrite, halite and kaolinite.

• The accessory minerals are illite, jarosite, gypsum, carbonates (such as, calcite, kutnohorite, Mn calcite, magnesite), and traces of others (alunite, anatase, anhydrite, aphtitalite, bassanite, ferrihydrite, fibroferrite, gibbsite, illite-smectite, nesquehonite and smectite).

• The key clay minerals are kaolinite, illite and smectite (in order of abundance).

• Pond 7 contains slightly more pyrite than Pond 10 (Appendix 3).

• The pond dyke walls are a source of ASS leachate to the pond.

• The sediments at the base of the pond are largely reduced (PASS) but the introduction of oxygen will create ASS and the associated ASS leachate.

• The water table is shallow under Pond 7, and at high tides influences the water chemistry in Pond 7 by estuarine/groundwater seepage.

• Seawater is a strong acid buffer, and the height of the water table is assisting in neutralising some of the ASS leachate in the Tomei pond sediments.

8.8 Summary

In the Pond 7 sediment core data, when there is a high percentage of clay (compared to other intervals in the core), there is also a greater percentage of organic matter (LOI% analysis); sediment pH is the lowest in the core when the S concentration in the core is the highest (from LECO analysis). Pond 10 cores follow the same trend. Pond 10 has a larger concentration of carbon in the

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Chapter 8: Pond Sediments sediment cores than was found in the Pond 7 cores. Each core contains a smaller concentration of S in Pond 10 compared to the sediment in Pond 7, but there is no discernable difference in the effect on the sediment pore water pH. This may be because Pond 7 is estuarine groundwater buffered.

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9 POND HYDROCHEMISTRY – PHYSICAL VARIABLES

9.1 Introduction

Physical variables are the measured parameters that will change rapidly after sampling, and therefore must be measured at the sample point to ensure optimal hydrochemical data quality. The following physical variables were collected by the author during field sampling: pH, water temperature, DO, EC, and Eh. Various aspects of the hydrochemical data from Ponds 7 and 10 are described in the next few chapters and referred to in Appendix 4. It is important to note that the following text presents data collected from Ponds 7 and 10 but a direct comparison between the two ponds should not be made due to their physical and chemical differences. It is also important to note that the author has referred to Eh measured in milli volts throughout the thesis. To convert Eh readings made in the field to true Eh values (the standard hydrogen electrode or SHE reference), one must add 244mV to the field reading. This conversion should be made to all reported Eh values in the thesis when comparing to the SHE.

9.1.1 pH

It is important to accurately measure the pH (or otherwise known as the hydrogen ion activity) when sampling water. In-situ measurement and recording of pH provides the only representative reading of the H+ ion concentration in the water. pH is measured on a scale of 1-14; 7 being neutral; 1-7 being acidic, and 7-14 being alkaline.

Hydrochemical and biochemical reactions occurring in the water column are influenced by pH. By observing the variation in pH throughout a geochemical system it is possible to establish the likely chemical inter-relationships between the systems organisms, sediments and the water column.

The optimal pH range for P. monodon is 7.5-8.5 (Chien, 1992). A similar pH is likely to be optimal for P. japonicus, however published work is not available to confirm this.

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9.1.1.1 Pond 7

In Pond 7 there is a notable difference between the measured pH in the water column and that of the pore water (Figure 32). The lowest (most acidic) pH values measured are in the sediment pore water samples (Table 16 and Figure 33a).

Figure 32: pH versus Depth from the Pond 7 water column and pore water samples over the three sampling periods

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Table 16: Range of recorded Physical Variables recorded for Pond 7

Variables Date Water column Pore water Sand ASS Max Min Max Min Max Min Max Min pH (units) Nov-01 8.18 7.98 7.92 6.79 7.92 6.90 7.48 6.79 Feb-02 8.246.97 7.56 6.51 7.15 6.51 7.56 6.62 Apr-02 8.177.56 7.64 7.26 7.64 7.26 7.64 7.26 Temp (°C) Nov-01 30.6 24.7 29.6 24.5 28.0 24.8 29.6 24.5 Feb-02 38.0 27.8 35.5 30.3 35.4 32.0 35.5 30.3 Apr-02 31.2 23.2 29.7 21.7 27.8 21.8 29.7 21.7 DO (mg/L) Nov-01 10.36 2.82 6.78 1.19 6.78 1.19 2.88 1.57 Feb-02 5.000.76 3.54 0.51 1.74 0.51 3.54 0.52 Apr-02 7.490.27 3.01 0.04 3.01 0.04 2.00 0.20 EC Nov-01 52,350 50,250 52,875 51,500 52,875 51,500 52,525 51,550 (μS/cm) Feb-02 53,125 52,400 53,525 50,675 53,525 52,700 51,325 50,675 Apr-02 51,300 50,725 52,000 50,950 51,750 50,950 52,000 51,375 Eh (mV) Nov-01 -53.0 -142.0 -110.5 -171.0 -115.0 -171.0 -110.5 -161.5 Feb-02 53.0 -352.5 -162.5 -378.5 -217.0 -378.5 -162.5 -373.5 Apr-02 249.6 -322.1 1.0 -319.7 1.0 -308.4 0.0 -319.7

Figure 33: TDS versus pH - yellow shading shows the optimal pH range for the P. japonicus; (a) Pond 7 and (b) Pond 10 – note that the lowest pH values are in the ASS pore water and sand layer for both ponds and that the more acidic samples are typically more saline

The pH of samples collected from the water column are within the optimal range (7.5-8.5) defined by Chien (1992). As P. japonicus are sediment dwelling species it is important to concentrate on the pH of that environment: the sediment as measured by the sediment pore water samples. The pH of the sediment pore water in Pond 7 over the growout season ranged from 6.5 to 7.9; lower than the preferred range offered by Chien (1992). Chien (1992) states that moderately low pH can have an adverse effect on prawn growth;

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Chapter 9: Pond Hydrochemistry – Physical Variables whereas extremely low pH (<6) can stress the prawn, cause soft shell, and can increase prawn mortality. The sand/sediment dwelling prawns in Pond 7 lived in sub-optimal pH conditions and the pond production data (see Section 12.5) demonstrates that they were adversely affected.

The gridding and contouring of chemical transects through the pond (see the plan view schematic of the transect in Figure 28) provides a powerful tool for summarising hydrochemical variation. Using temporal analysis to look at changes in pH through the pond transects over the growout season has proved a valuable interpretation and presentation tool.

Figure 34 focuses on pH variation in Pond 7. In November, acidic water was largely observed in the underlying ASS and sand pore water from the centre of the pond. In February, the acidity spread throughout the sediment and decreased the pH of the water column. In April, the acidic waters are restricted to the centre of the pond.

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Figure 34: Time series of contoured pH transects for Pond 7 with 0cm representing the sand-water interface

Prior to review of the hydrochemical data it was postulated that maximum acidity from the ASS in these ponds should be at the beginning of the season when the ponds are initially filled. The logic being that during pre-season pond preparation, the pond was empty and the underlying ASS was exposed to four months of rainfall and atmospheric oxygen. This exposure should result in ASS

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Chapter 9: Pond Hydrochemistry – Physical Variables oxidation, and the generation of an acid leachate that would dominate early pond chemistry (November sampling).

This was not observed in Pond 7, the most acidic samples were measured in February (Figure 32), and the author postulates that this is due to: (1) the farmers practice of liming the pond during preparation (this was effective in neutralising the ASS leachate at the beginning of the season); (2) seawater initially buffered the ASS leachate. The February acidity is interpreted to show that with time continued ASS oxidation generated leachate which eventually neutralised all of the lime and then exceeded the buffering capacity of the seawater.

The time series of contoured cross sections of pond and pore-water pH for Pond 7 presented in Figure 34 demonstrate that acidity was closely related to the pond dyke walls and base (the ASS), with the lowest pH in each piesometer being recorded in the sediment water samples. As the grow-out season progressed, this time series shows that as the system moved towards equilibrium as acidity migrated from the pore water into the water column, effectively resulting in the pond sediments becoming more alkaline.

The two piesometers that recorded the lowest pore water pH values during all three sampling rounds were located closest to the dyke walls. It is apparent that the dyke walls generated and contributed acidic runoff to Pond 7. This is in good agreement with the sediment core data presented in Chapter 8.

9.1.1.2 Pond 10

As in Pond 7, there is a notable difference between the measured pH in the water column and the pore water in Pond 10. The lowest (most acidic) pH values for all three sampling rounds were measured in the sediment pore water (Figure 35 and Table 17).

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Figure 35: Depth versus pH for the Pond 10 water column and pore water samples over four sampling periods Table 17: Range of recorded Physical Variables in Pond 10

Variables Date Water column Pore water Sand ASS Max Min Max Min Max Min Max Min pH (units) Nov-01 8.71 8.20 8.01 6.78 8.01 6.95 7.72 6.78 Feb-02 8.638.29 7.68 6.88 7.30 6.88 7.68 7.48 Apr-02 8.648.32 8.15 6.86 8.15 6.86 7.79 7.35 Jun-02 8.998.51 8.28 7.16 8.28 7.20 7.39 7.16 Temp (°C) Nov-01 30.7 23.4 33.2 23.8 30.9 25.0 33.2 23.8 Feb-02 39.1 27.5 40.0 26.4 35.0 29.8 40.0 26.4 Apr-02 29.5 20.0 28.7 22.0 27.5 22.7 28.7 22.0 Jun-02 20.8 15.2 21.8 18.3 20.4 18.3 21.8 19.6 DO (mg/L) Nov-01 3.81 2.43 1.62 0.43 1.18 0.43 1.62 0.68 Feb-02 6.002.41 1.62 0.14 1.46 0.14 1.62 0.17 Apr-02 5.653.34 1.51 0.02 1.51 0.02 1.44 0.30 Jun-02 6.314.37 1.76 0.67 1.76 0.67 1.49 0.88 EC (μS/cm) Nov-01 51,500 50,975 52,375 51,200 52,375 51,325 51,725 51,200 Feb-02 54,300 53,350 54,050 51,400 54,050 53,600 53,075 51,400 Apr-02 53,325 52,550 53,900 52,700 53,900 52,700 53,400 52,750 Jun-02 49,175 48,825 49,250 48,800 49,150 48,800 49,250 48,825 Eh (mV) Nov-01 -79.5 -119.5 -108.0 -248.5 -136.0 -248.5 -108.0 -153.5 Feb-02 -172.0 -209.0 -245.0 -377.5 -286.5 -356.0 -245.0 -377.5 Apr-02 42.8 -107.9 -52.2 -333.6 -90.2 -322.0 -52.2 -333.6 Jun-02 360.1 8.9 -11.1 -209.7 -11.1 -209.7 -33.4 -160.4

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The pH of the water column in Pond 10 ranges between 8.2-9.0 and is therefore notably higher (more alkaline) than was observed in Pond 7. Similarly the pH of the sand and ASS pore water is slightly higher than that in Pond 7. This is probably due to: (1) the sediments in Pond 10 being less dominated by PASS; and, (2) Pond 10 being volumetrically larger and therefore (by ratio) the seawater has better buffering capacity.

The pore water in Pond 10 was the most acidic in November; however the pond water has a narrower pH range through the growout season than that of Pond 7.

The minimum pH in the sand pore water in November was 6.95, however only 200 mm above this sample point (in the water column) a pH of 8.20 was measured (Figure 35). This steep pH gradient must have stressed the prawns as they moved between the sand layer and the water column.

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Figure 36: Time series of contoured pH transects for Pond 10 with 0cm representing the sand-water interface

Throughout the growout season the centre of the pond was discharging acidic leachate to the pond. To minimise stress the prawns are likely to have moved to the less acidic (near dyke wall) portions of the pond. This relocation would have increased population density, competition, and resulted in reduced pond efficiency and productivity.

9.1.1.2.1 Comments relating to pond pH

Boyd (1992) stated that “liming shrimp ponds during the grow-out period is seldom necessary because brackish water is well-buffered” (pg 166). This view is widely held in both the scientific community and the Australian prawn industry with regard to ASS, however the transects and cross-plots presented above show that the seawater buffering capacity is unable to fully counterbalance the acidity released by ASS. Therefore at the study farm (and at other PASS/ASS farms) it is important to monitor and regulate pond pH and avoid an over reliance on the exchanged seawater to buffer any generated acid.

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It is also essential to recognise that the P. japonicus species reside in the sand layer and feed in the water column. When there is a significant geochemical gradient between these two environments the prawn will be stressed when it moves from one area to the other.

Incorporating pond pH management systems into farm practices will aid in building sustained pond productivity and a more reliable income stream.

9.1.2 Water Temperature

Water temperature is an important physical variable for geochemical modelling as it is required for calculating chemical equilibrium and mineral solubilities. The method of field collection used in this study was not ideal: the author recognises that there is likelihood that the temperature data is overstated. This is due to high ambient temperatures and thermal equilibration during sample collection (the samples were pumped from the sample location through non- insulated plastic tubes to the sample collection followed by temperature measurement and sampling point).

9.1.2.1 Pond 7

The preferred temperature range for P. japonicus is 24-29°C (Coman et al., 2002). The maximum recorded water temperature was measured in February (38.0°C in the water column, and 35.5°C in the pore water), and the minimum temperature was measured in April (23.2°C in the water column, and 21.7°C in the pore water): see Figure 37.

These recorded temperatures are close to ambient temperatures during sampling (see the ambient temperature data presented in Chapter 2). If these data are representative of in-situ pore water temperatures the Pond 7 data indicate that the P. japonicus were stressed by high temperature during February.

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Figure 37: TDS versus Temperature with the P. japonicus comfort zone shown by the yellow shading for (a) Pond 7 and (b) Pond 10 – note the relationship between higher temperatures and higher TDS

Ponds are classified as stratified when the pond temperature difference between the top and the base of the pond is >1.0 °C and well mixed if the difference is <0.5 °C (Costa-Pierce and Laws, 1985). Data presented in Table 17 show that there is a 7-12 °C temperature difference depending on which sampling round is being considered. Even though mechanical mixing devices (paddlewheels and aerators) were being actively used, pond 7 is thermally stratified (Figure 38).

Figure 38: Pond 7 water temperature (oC) over the sampling period showing a 7-12oC temperature variation in the pond depending on which sampling round is considered

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9.1.2.2 Pond 10

The lowest water temperatures were recorded in June (winter) in both the water column and sediment pore water samples. The highest temperatures were measured in the February samples with a maximum recorded temperature of 40°C in both the water column and the pore water. Figure 39 shows that like Pond 7, Pond 10 is thermally stratified.

Figure 39: Pond 10 water temperature (oC) over the sampling period showing a 7-15oC temperature variation in the pond depending on which sampling round is considered

If these data are representative of the in-situ water temperature at the time of sampling then data from Pond 10 indicates that the P. japonicus were affected by high temperature during February and low temperature in June.

9.1.3 Dissolved Oxygen

Dissolved oxygen (DO) is measure of the available free oxygen in water and concentrations are measured in milligrams of oxygen per litre of water (mg/L). The concentration of oxygen is important to all respiring organisms. The DO concentration also controls reduction-oxidation (redox) reaction rates as organisms are forced to use alternate electron donors to drive their metabolic processes.

P. monodon is best suited to DO concentrations above 3 mg/L (Chien, 1992). P. japonicus is thought to be able to survive in lower concentrations of DO than

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P. monodon (Chien, 1992) however an optimal DO concentration for P. japonicus was not located in published literature.

9.1.3.1 Pond 7

Table 16 and Figure 40 show that the maximum DO concentrations in Pond 7 are 10.36 mg/L (in the water column), 6.78 mg/L (in the sand) and 3.54 mg/L (in the ASS). All these concentrations are high enough for P. japonicus to live relatively comfortably. However, the minimum recorded concentrations in the pond are 0.27 mg/L in the water column, 0.04 mg/L in the sand and 0.20 mg/L, in the ASS (Figure 41). These concentrations are well below optimal concentrations and will lead to hypoxia and death of the aquaculture species.

Figure 40: Dissolved oxygen in (a) Pond 7 and (b) Pond 10 over time

Figure 41: TDS versus DO for (a) Pond 7 and (b) Pond 10 (yellow shading is the zone of hypoxia for P. japonicus; note the same symbols are used in all cross plots, so refer to other graphs for a legend)

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November had the highest DO concentrations out of the three sampling periods, with April having the lowest (Table 16). It is logical that the highest concentration of DO was measured in the pond at the beginning of the growout season. The pond remained empty for four months before it was filled. This meant that the pond was exposed to atmospheric oxygen for four months before the first collected sample in November; there was little time for organic matter to build-up in the pond; only a partial establishment of the ponds biological community; and, only juvenile aquaculture species in the pond: all resulting in a low oxygen demand which therefore had minimal impact on DO concentrations.

Over time, as the biota in the pond proliferated and the aquaculture species grew, the ponds oxygen demand increased and DO progressively decreased. With this change in dissolved oxygen availability (both at the sediment-water interface and in the sediments) anaerobic organisms progressively replaced their aerobic predecessors (sulfur anaerobes proliferate under reduced conditions where hydrogen and sulfur are contributed by the oxidation of ASS).

9.1.3.2 Pond 10

The lowest average concentration of DO in the water column (2.98 mg/L) was measured in November (see Table 18). There is a relatively large difference between the average water column measurements in November and average data from February and April. This may be due to poor paddle wheel placement, low efficiency or non-use, or it may demonstrate an increase in the pond oxygen demand due to episodic algal bloom(s).

The average concentrations of DO in the sand layer (where the prawns live) is around 1 mg/L. Egusa (1961) showed that when DO falls to this concentration, prawns move to an alternative location in the pond where there is more available oxygen (in the case of Pond 10, the water column) and this would: (1) stress the prawn; and, (2) force it to live in the water column. Living in the water column is unnatural for P. japonicus and exposes them to predators and solar radiation.

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Table 18 shows there was a progressive increase of average DO in the water column through the growout season and that the concentration of DO was low in the pond sediments (sand and pore water) in February.

Table 18: Average Dissolved Oxygen (DO in mg/L) concentrations for Pond 10

Date Variable Average water column Average sand Average pore water Nov-01 DO 2.98 0.85 1.07 Feb-02 DO 4.04 0.53 0.63 Apr-02 DO 4.30 0.72 0.77 Jun-02 DO 5.01 1.08 1.12

In general, DO in the sand and pore water is higher in the cooler months, and this is indicative of reduced biological activity (lower temperatures, less solar radiation and shorter days).

9.1.4 Eh (Redox)

Eh (or redox potential) is measured in millivolts (mV) with values above zero containing free oxygen in solution, and values below zero showing there is no dissolved oxygen in solution (the water is anoxic).

Figure 42: TDS versus Eh for (a) Pond 7 and (b) Pond 10 – note that lower (more negative) measured Eh tends to accompany higher water salinity

In both Ponds 7 and 10 there is a tendency for increased TDS with lower (more negative) Eh (Figure 42). pH also has a direct relationship with Eh; when the pH is low, reduced elements are dominant and the Eh decreases (Figure 43). The lowest Eh measurements are measured in the sand and pore water samples - the areas where there is poor oxygen availability and high

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Chapter 9: Pond Hydrochemistry – Physical Variables consumption through the oxidation of ASS (Figure 44). The presence of the ASS in these ponds is interpreted to be the primary reason for this relationship.

Figure 43: pH versus Eh for (a) Pond 7 and (b) Pond 10 – note that lower measured Eh tends to accompany low (acidic) pH measurements

Since Eh is related to the presence of oxygen, it should correlate with measured DO concentrations. Eh changes rapidly when water comes in contact with the atmosphere or an alternate O2 source, so taking reliable Eh measurements is difficult (Fetter, 1988). Inspection of Figure 44 results in the conclusion that very few samples have a positive Eh, even though all samples were recorded to contain DO. This is not intuitively correct; accordingly the Eh values are used as a relative - not absolute - measure for geochemical interpretation.

Figure 44: Eh versus DO for (a) Pond 7 and (b) Pond 10 (see other bivariate plots for legend)

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Berner (1981) suggested that electrode measured field Eh will differ to calculated Eh (produced by hydrochemical modelling software) because the platinum and gold electrodes identify the dominant “electroactive” redox species in the system. The electroactive species may not be the redox couple that is dominating this aqueous system at the time of measurement. Therefore, Berner (1981) suggests that electrode-based Eh field measurements should be used cautiously, and instead hydrochemical modelling software (redox couple based) should be used to calculate the water samples Eh (usually as pE).

Figure 45: pE versus Eh diagram for pore water samples in Pond 10

Figure 45 shows measured Eh versus calculated pE and demonstrates that there is not a good match between these variables. The notable feature of this graph is that the highest calculated pE (and lowest measured Eh) values are all related to the sediment derived water samples. Clearly the bulk of the anaerobic geochemical reactions are occurring in the pond sediments where there is little or no available oxygen.

Figure 46 shows Eh data from Pond 10 verses time. It shows that the most reduced samples (lowest Eh) were collected from the pore water in the warmer months at the beginning of the growout cycle. It is worthy of note that Pond 10 became more oxidised in both the sediment and water column as the season progressed.

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Figure 46: Eh measurements from Pond 10 over the sampling period

In both Ponds 7 and 10 the pond water column contains a larger concentration of dissolved oxygen than the sand or clay pore water samples. The comfort zone for the P. japonicus prawn cuts off when DO falls below 2 mg/L: in the Tomei ponds this generally at the top of the sand layer (water column-sand interface). Consistent with the lower DO concentrations, the measured Eh for the sand and pore water samples are the lowest in the dataset (being generally below -200mV).

The Eh data supports the DO data: neither the sand nor the ASS pond base or dyke walls provide a suitably oxygenated residence for the P. japonicus prawn.

9.1.5 Electrical Conductivity (Salinity)

Electric conductivity (EC) is a measure of the ability of a fluid to conduct an electric current. EC is typically measured in micro siemens (μS) per cm2. This conductivity is dependent on both the salinity and chemical composition of the fluid: EC is a good first-pass guide to estimating salinity.

Figure 47 shows the relationship between TDS and EC with the calculated line of best fit: the goodness of fit (R2) is much better for Pond 10 than it is for Pond 7.

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Figure 47: EC versus TDS showing the relationship between EC and salinity for the Tomei waters (a) Pond 7 and (b) Pond 10

P. monodon is able to tolerate salinities ranging from 3,000-71,000 mg/L (Solis, 1988). The optimal concentration of salinity largely depends in the size of the animal. It is thought that juveniles do better at lower salinities (~25,000 mg/L) and adults at higher salinities (~35,000 mg/L). As with P. monodon, P. japonicus adults grow well at 35,000 (approximately 50,000μS) (Arnold, 2005 pers. comm.).

9.1.5.1 Pond 7

Table 16 shows that all water samples have the highest recorded EC in February 2002. This is because pond evaporation is greatest in the summer (Figure 48); February is also the month with the lowest rainfall (Figure 6) so there is minimal dilution of the accumulated salts in the pond or adjacent estuary. During this month the Pond 7 water level was highest relative to the pond dyke walls (when compared to the other monitoring months) making the pond deeper and reducing the efficiency of the paddle wheels (reduced mixing and aeration).

During February farm management exchanged pond water with the adjacent estuary more often (in an attempt to manage the increasing salinity). These efforts had little impact on pond salinity as the estuary intake waters were equally saline (as a result of low rainfall, intense evaporation, poor circulation within the estuary, and limited access/exchange with the less saline Pacific Ocean).

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Figure 48: EC versus Cl- for (a) Pond 7 and (b) Pond 10 (see other bivariate plots for legend)

9.1.5.2 Pond 10

Like Pond 7, Pond 10 had its highest EC recorded in the February, and lowest EC in June (Table 17).

In general the EC is similar in the water column, sand and ASS pore water (suggesting hydraulic connectivity). However, in February there is a slight drop in Pond 10 EC across the interface between the sand layer and the ASS pore waters (lower EC in the ASS). This was not observed to the same extent in Pond 7. In Pond 10 this is possibly attributed to there being a lower hydraulic conductivity of the ASS in this month (clay only lets water pass through very slowly when compared to sand). This low hydraulic conductivity could mean that the February spike in the pond water EC did not have sufficient residence time in the pond for the establishment of an EC equilibrium between the pond and the ASS pore waters.

9.2 Summary

The physical variables measured from the Tomei farm Ponds 7 and 10 indicate that the growing environment is not ideally suited to the sediment dwelling P. japonicus.

The reasons are summarised as:

• pH is too low (acidic), and too variable (significant pH gradients being present in the ponds)

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• Water temperature is too high in summer, and too low in winter

• DO (or Eh) is too low in the sand (where this species reside) to support comfortable respiration

• EC (salinity) is very high in February and this spike indicates poor exchange between the farm intake/adjacent estuary system and the ocean.

Overall it is not surprising that the prawn productivity from the Tomei ponds was low in 2001-2002 as the physical variable data show that the pond conditions were not suited to the aquaculture species. These P. japonicus prawns were most likely often stressed as supported by the reported high mortality rates and low overall pond productivity.

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients

10 POND HYDROCHEMISTRY - MAJOR IONS & NUTRIENTS

10.1 Introduction

Chemical analysis for major ions were undertaken using various analytical techniques. For the purpose of this summary the ions are categorised into two + + 2+ 2+ - 2- - groups; the cations (Na , K , Ca , Mg ) and the anions (Cl , SO4 , HCO3 ).

Bivariate plots are used through the chapter to depict variations in water sample chemistry relative to five reference seawater samples. These seawater samples are used to illustrate differences between the chemistry of the water in the Tomei ponds and their primary (pre-pond entry) composition. Establishing a representative seawater composition is critical because these reference compositions provide the basis for the interpretation of the active geochemical reactions in the pond(s).

The order of abundance of cations and anions for both Ponds 7 and 10 are the same, and are as follows: Na>Mg>Ca>K and Cl>SO4>HCO3. The dominant ions are Na and Cl.

10.2 Establishing a representative seawater composition

The worlds’ oceans are not chemically homogenous. Seawater has variable chemistry and salinity depending on the local environmental conditions. This variation is created by both natural sources and anthropogenic activity; some examples are currents, tides, wind, or input from rainfall, rivers, groundwater discharge and agricultural, domestic or industrial runoff or losses from evaporation and the associated precipitation of minerals. This natural variability in seawater can be reduced to two fundamental independent variables (salinity and chemical composition). The presence of these two independent variables requires that a reference line (rather than a reference point) be used for hydrochemical analysis. Since seawater evolves initially from rainwater, the intercept for this seawater line (of best fit) on all graphs is zero (rain water has essentially no dissolved salts).

The author’s representative seawater line uses both existing and new chemical data. A literature review provided two published seawater analyses (Hem,

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1989; and http://ozreef.org/reference/composition.html); and the collection and analysis of three samples from the Pimpama estuary system. The location of these three samples is shown in Figure 2 and the associated data is referred to in Appendix 5. As a component of the quality control methodology, duplicate samples were taken for each of the Pimpama estuary samples, and these were put through the full suite of analyses in conjunction with the primary samples.

All samples analysed in this study have seawater salinity and this requires specific analytical methodology be used on the ICP-OES instrument. Samples were diluted by a factor of 10 (1 part sample: 10 parts millicule water). Normal measurement error in the manual dilution of these samples has introduced unavoidable analytical error.

In an attempt to remove outliers and more obvious spurious results, the chemical analysis of each sample has been validated by calculating the Charge Balance Error (CBE) (see Equation 16). If samples fall within ±5% of zero CBE they are accepted as being analytically valid and were incorporated into the interpretation (Figure 49), if the sample failed this test the analysis was repeated: if the repeat testing still resulted in an unacceptable CBE the sample was not used in the interpretation.

Figure 49: Summary of CBE across the data set for (a) Pond 7 and (b) Pond 10

The CBE calculations of the collected seawater reference samples indicate that the Tomei intake and estuary waters were acceptable. There is however differences in concentrations and ratios between each of the primary and

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients duplicate samples. This is put down to analytical error and so an average ionic concentration was calculated (based on the primary and duplicate sample pairs) to provide the most representative seawater end-member sample. The Jumpinpin Bar duplicate failed the CBE and was ignored. The primary Jumpinpin Bar sample (on a CBE basis) was marginally acceptable at +4.4%, but after comparison with the available seawater data this sample was deemed unreliable, and was not used in the interpretation provided below.

The authors’ representative seawater composition line is based on the following samples: (1) Great Barrier Reef (GBR) Seawater (http://ozreef.org/reference/composition.html); (2) Hem Seawater (Hem, 1989); (3) averaged Tomei intake water, and; (4) averaged Tomei estuary water.

On each graph this “seawater line” is shown as a solid pink line.

10.3 Total Dissolved Solids

Total Dissolved Solids (TDS) is a chemical measure of water salinity. It is the mass of salt (in milligrams (mg) or grams (g)) that is precipitated when you evaporate one litre of water.

The TDS of both Pond 7 and 10 varies throughout the growing season from about 33,000 to 45,000 mg/L (33-45 g/L). Average seawater salinity is reported by Hem (1989) to be 35,000 mg/L: most of Pond 7 and 10 chemical data have salinities that are higher than this average (Figure 50).

Figure 50: Calculated TDS versus Electrical Conductivity for (a) Pond 7 and, (b) Pond 10

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Over the season there is 20% variation in pond salinity: better managing this fluctuation is likely to reduce species stress and improve pond productivity.

Pond salinity is controlled by: (1) the salinity of the intake water; (2) the amount of pond water evaporated; (3) the amount of rainfall and runoff entering ponds, and (4) the volume and frequency of water exchanges (both anthropogenic and natural).

In both Ponds 7 and 10 there is a correlation between pond salinity and climate during the grow-out season: salinity is moderate in November 2001, high in February 2002, moderate in April 2002, then where data is available from pond 10 low in June 2002 (Figure 51).

Figure 51: Cl- versus TDS for (a) Pond 7 and, (b) Pond 10

This variation in salinity is attributed to evaporative concentration and/or rainfall as these directly influence the in-situ pond salinity, and control the salinity of the estuary system (and therefore the intake water). Note that Cl- is a conservative element in water: it accumulates without being effected by chemical reactions; its concentration only decreases if there is fresh water mixing which has a dilution effect, or when chloride salts precipitate from a super-saline fluid.

10.4 Saturation Indices

Saturation Indices (SI) are thermodynamically calculated parameters that indicate whether a water is chemically supersaturated (SI>+0.5), in equilibrium (+0.5>SI>-0.5), or undersaturated (SI<-0.5) with respect to a mineral phase.

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If the saturation indices suggest that the water is supersaturated (SI>+0.5) with respect to a specific mineral, then that mineral has the chemical potential to precipitate from solution, but will not dissolve into the solution: this does not however imply that this reaction is active in the system. In the same way, if the mineral is undersaturated with respect to the solution, it can dissolve into that solution, but will not be precipitated from the solution.

Saturation indices are important as they indicate the thermodynamically calculated potential for a mineral to react (i.e. dissolve or precipitate), they are not proof that the mineral is precipitating or dissolving in that particular solution. The SI is a tool for geochemical interpretation to ensure that any postulated mineral dissolution or precipitation reactions are thermodynamically feasible (given the chemical state of the solution).

PhreeqCI (Parkhurst and Appelo, 1999) is a freeware windows computer based hydrochemical modelling programme. This programme was used to calculate the SI’s for the water samples from each of the sampling rounds in Ponds 7 and 10 (the SI’s referred to in Appendix 6).

10.5 Sodium (Na+)

Sodium can accumulate in water by: (1) the dissolution of Na-rich evaporite minerals (chloride, carbonate or sulfate); (2) ion-exchange reactions where Na+ is released to solution and Ca2+, Mg2+ or K+ are taken up; (3) desorption releasing Na+ from clay surfaces in response to environmental chemical change (in the presence of high cation-exchange capacity clay); (4) the weathering of Na-rich silicate minerals such as albite (NaAlSi3O8); (5) ocean- derived salty aerosols; (6) dissolution of pond preparation products such as Na-hypochlorite (NaOCl); or, (7) sea-salts carried in coastal rainwater

Sodium can be removed from water by: (1) ion-exchange reactions taking up 2Na+ and releasing Ca2+, Mg2+ or 2K+ to the water; (2) adsorption on the surface of clay minerals; and, (3) the precipitation of Na-rich evaporite minerals during the evaporation of water.

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10.5.1 Observed Sodium concentrations and potential source/sink

Figure 52 shows that there is generally an excess of Na+ in both ponds when compared to the reference seawater line.

Figure 52: Na+ versus Cl- (meq/L) for (a) Pond 7 and, (b) Pond 10

Sodium is greater than the reference line by 0-30 meq/L in both ponds. This is possibly attributed to: (1) Na-hypocholrite dissolution (NaOCl); (2) 2Na+Ca2+ ion-exchange; (3) desorption from clay surfaces; and/or, (4) sodium-rich evaporite (carbonate or sulfate) dissolution. Each of these potential source reactions are described below.

(1) Na-hypocholrite (NaOCl) is spread (as a pond sterilisation agent) over the pond base and dyke walls prior to the start of the grow-out season. This agent dissolves to produce Na+ and Cl- ions in a ratio of 1:1. As there is a greater amount of Na+ than Cl- (not a 1:1 relationship) indicating that this is not the source of the excess Na+.

(2) The release of Na+ through 2Na+Ca2+ ion-exchange can only explain a small portion (about 10%) of the ion-exchange: there is approximately a 20 meq/L excess of Na+, and only a 2 meq/L short fall in Ca2+ (see Section 10.6). This process is probably contributing to the excess Na+, but is not the dominant process.

(3) Desorption leads to a potential increase in Na+ through the release from Na-loaded cation-exchange sites on clay surfaces. This occurs in response to a change in chemical potential in the system (i.e. when new equilibrium

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients conditions develop in the pond: an example being water exchange events where the intake water has a different chemistry to the pre-existing pond water). This process is likely to be actively contributing Na+ to this system, but this contribution can not be tested with the available dataset.

(4) Evaporite minerals are present in the pond sediments at the start of the grow-out season (derived from evaporation of tidally influenced groundwater, or the transport of saline aerosols when the pond was being prepared for the new season and was empty).

The dissolution of halite (NaCl) is not the Na+ source because this reaction releases Na and Cl in a ratio of 1:1. Figure 52 demonstrates that the excess sodium is derived from a reaction that does not contribute chloride ions.

The salts are potentially sulfate and/or carbonate salts as the dissolution of Na-carbonate or Na-sulfate will not impact the Cl concentration in the pond. The saturation indices (Appendix 6) indicate that Thenardite (a Na-sulfate mineral) is able to dissolve; similarly Themonarite, Trona, Natron (Na- carbonate minerals) are able to dissolve. There is insufficient mineralogical data available to confirm that these minerals are present in this system but this is a good potential Na+ source for this system.

Figure 52b shows that Na+ is up to 30 meq/L enriched when compared to the reference seawater line. There are however notable differences between the sampling rounds in Pond 10. In February, April and June 2002 the data is spread either side of the seawater line. The Na+ shortfall in these samples is best explained by Na-adsorption onto pond sediments (the reverse reaction to that described in point 3 above).

10.5.2 Sodium - spatial distribution in the ponds

Figure 53 presents Na concentration transects for Pond 7 during November, February and April: the ponds are poorly mixed (heterogeneous). In November, the highest Na+ concentrations were sampled in the pore waters. In February, Na+ is highest in the pond water (as for all ions) due to evaporative concentration. In April, the pond mixing is ineffective (greatest Na concentration gradient is across the profile in the water column) and there

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients appears to be a groundwater-related/estuarine tidal influx introducing a Na-rich slug through the pond base in the deeper (centre) portion of the pond.

Figure 53: Time series of contoured sodium concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02)

Pond 10 is relatively homogeneous at every sampling period throughout the grow-out season (Figure 54). In November, the Na+ concentrations are relatively homogeneous between the water column and the pore water samples. The slightly lower concentrations of Na+ are located nearer the pond walls, suggesting that there are pockets of lower concentration water circulating in the pond that are not mixing. In February and April, salinities are highest due to evaporative concentration.

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It is interesting that in Pond 10 the Na+ concentration during all sampling periods is relatively similar in the water column and pore water. This suggests that the pond base and dyke walls in Pond 10 could exchange water (that is, they have sufficient hydraulic conductivity to exchange water which results in equilibration/removal of concentration gradients). In June, salinities are lowest of all the sampling periods and most Na+ is associated with the dyke walls.

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Figure 54: Time series of contoured sodium concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)

10.6 Calcium (Ca2+)

Calcium can accumulate in pond waters by: (1) the dissolution of Ca-rich evaporite minerals (carbonate or sulfate); (2) ion-exchange where Ca2+ is

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients released to solution as Mg2+, 2Na+ or 2K+ are taken up; (3) desorption releasing Ca2+ from clay surfaces in response to environmental chemical change (high cation-exchange capacity clay); (4) the weathering of Ca-rich silicate minerals; (5) ocean-derived salty aerosols; or, (6) sea-salts carried in coastal rainwater

Calcium can be removed from these waters by: (1) ion-exchange reactions taking up Ca2+ and releasing Mg2+, Na+ or K+ to the water; (2) adsorption on the surface of clay minerals; and, (3) the precipitation of Ca-rich evaporite minerals during the evaporation of water; or, (4) biological uptake – either plants or animals for nutritional, skeletal or calcareous development.

10.6.1 Observed Calcium concentrations and potential source/sink

It is important to note that the ponds are limed at the beginning of the grow-out season in an attempt to counteract the acid being generated by the underlying ASS.

2+ Figure 55 shows that both Ponds 7 and 10 are depleted in Ca when compared to the seawater reference line. This suggests that the Ca2+ is being removed from the pond water by chemical and/or biological mechanisms.

Figure 55: TDS versus Ca2+ (mg/L) for (a) Pond 7 and (b) Pond 10

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Figure 56: Cl versus Ca2+ (meq/L) for (a) Pond 7 and (b) Pond 10

This shortfall in Ca2+ (Figure 56) as discussed in Section 10.5 is partly attributed to ion exchange with sodium in the ratio of 1:2. It is also possible that the Ca2+ was lost through CaMg ion exchange, but this only holds true for November 2001.

There is also likely to be Ca2+ loss from solution through Ca-rich mineral precipitation (a mineral such as calcite (CaCO3), aragonite (CaCO3) or dolomite (CaMg(CO3)2) (which are all supersaturated).

• If dolomite was the precipitating carbonate then one would expect that Mg2+ would also be depleted in November and June. As it is not (it is actually in excess in November and lies on or around the seawater line in June) then dolomite precipitation is not responsible for the loss of Ca2+.

• Ca2+ might have been scavenged by aquatic organisms such as

polychaetes (worms) for use in building their CaCO3 burrows in the pond (Plate 42).

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Plate 42: CaCO3 worm structures on the floor of the pond bottom and showing iron oxide staining

The highest concentrations of Ca2+ occur in Pond 7 when the pond pH is lowest (most acidic) (see Figure 57; note that this is active in the pond sediments only).

At the start of the season (November 2001) calcium carbonate minerals (such as shell material or lime) are applied to the ponds to neutralise the ASS before the ponds are filled with water. The dissolution of this calcium carbonate-rich treatment of the pond base and dyke walls releases Ca2+ into the water column.

Figure 57: Pond 7 pH versus Ca2+ (mg/L) for (a) Pond 7 and, (b) Pond 10

In Figure 57b, the pH in Pond 10 does not have the same pronounced affect on the concentration of Ca2+ as seen in Pond 7. There is however a slight increase in the concentration of Ca2+ in Pond 10, particularly in November,

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February and April. This is probably due to Pond 7 being slightly more acidic than Pond 10.

Calcium concentrations in both ponds are being primarily controlled by: (1) Ca- carbonate dissolution under acidic conditions; (2) normal and reverse ion- exchange reactions; and (3) uptake by biota as they build exoskeletons.

10.6.2 Calcium - spatial distribution in the ponds

The greatest concentration of Ca2+ in Pond 7 was measured in the area associated with the pond dyke walls in November 2001 and February 2002 (Figure 58). This is likely due to liming affects and buried calcareous material.

In February and April 2002 the concentration of Ca2+ in the pond water and pore water is lower then it was at the start of the grow-out season. In both of these latter sampling periods, concentrations in the pond sediments are close to equilibrium with the seawater line, however the water column samples fall below this line.

Across all 3 sampling rounds there is a decrease in the measured concentrations of Ca2+. This progressive fall in Ca2+ concentrations is due to initial spiking associated with pre-season liming treatments followed by exhaustion/dissolution of the CaCO3 source through the grow-out season. The tightness of the April 2002 data shows most CaCO3 dissolution was complete prior to this sampling round.

Figure 58 shows that during February 2002 the dyke walls held the highest concentration of Ca2+. In April this switched with the pond centre. This is interpreted to be the result of either: (1) the progressive dissolution of Ca- carbonate treatment applied by farm management to counteract the pond productivity problems, and/or (2) the up welling of estuarine groundwaters that is inferred to have decreased pH and increased Na concentrations.

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Figure 58: Time series of contoured calcium concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02)

Out of the three sampling months, the highest mean Ca2+ concentration for Pond 7 is in the pore water from November 2001 (Table 19). In November (at the start of the grow-out season, once the initial liming ASS treatment had dissolved into the pond water), the pond was supersaturated with Ca2+ minerals such as calcite (CaCO3), gypsum (CaSO4 2H2O), anhydrite (CaSO4), aragonite (CaCO3) and dolomite (CaMg(CO3)2), and to a lesser extent portlandite (Ca(OH)2)). In the February and April 2002 sampling periods these minerals are undersaturated and therefore have the thermodynamic potential to dissolve.

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Preda and Cox (2004) noted that calcite (CaCO3) and anorthite (CaAl2Si2O8) is associated with the shale and sandstone deposits of the Pimpama flood plain.

Table 19: Mean concentrations of Ca2+ in the water column and pore water from Pond 7

Data Variable Average water column (mg/L) Average pore water (mg/L) Nov-01 Ca2+ 400.96 452.80 Feb-02 Ca2+ 452.69 437.80 Apr-02 Ca2+ 423.04 429.5

Figure 59 shows that in Pond 10 there is variation in the concentration of Ca2+ over the growout season which is similar to that of Pond 7 (i.e. the lowest concentrations of Ca2+ in Pond 10 were sampled towards the end of the growout season). However, unlike Pond 7, there are “hot spots” of Ca2+ towards the centre of the pond, particularly in November and February. This may be due to the dissolution of shell or other calcium (carbonate) material by acid leaching from the ASS which would have released Ca2+ into the adjacent sediment pore water.

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Figure 59: Time series of contoured calcium concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)

10.7 Potassium (K+)

Potassium can accumulate in the Tomei pond water by: (1) the dissolution of K-chloride; (2) ion-exchange - 2K+ is released into solution as Mg2+, Ca2+ or 2Na+ are taken up; (3) desorption releasing K+ from clay surfaces in response to environmental chemical change (high cation-exchange capacity clay); (4) the weathering of K-rich silicate minerals: feldspars (orthoclase and microcline), micas and feldspathoid leucite (KAlSi2O6) and the weathering of K- rich clays (Hem, 1989); (5) fertilisers used by Tomei or the adjacent farmers; (6) ocean-derived salty aerosols; or, (6) sea-salts carried in coastal rainwater

Potassium can be removed from water by: (1) ion-exchange reactions taking up 2K+ and releasing Mg2+, Ca2+ or 2Na+ to the water; (2) adsorption on the surface of clay minerals; and, (3) the precipitation of K-chloride; or, (4) biological uptake by plants for nutrition.

10.7.1 Observed Potassium concentrations and potential source/sink

Figure 60 presents the Potassium concentration in Pond 7 and Pond 10.

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Figure 60: TDS versus K+ for (a) Pond 7 and (b) Pond 10

Potassium is reacting conservatively without a natural geochemical source or sink (Figure 61).

Figure 61: Cl- versus K+ for (a) Pond 7 and (b) Pond 10

All water taken at Pimpama (including the pond water) contains a higher concentration of K+ than the reference water samples from the GBR and Hem (1989). The average K+ concentration in seawater (GBR and Hem) is about 390 mg/L. In Pond 7 the average concentration (over the three months of sampling) was 415 mg/L. This suggests that there is another source (other than naturally contained in seawater) contributing K+ to the pond water. The source of K+ in the pond water may be: (1) from the addition of fertilisers in the catchment; (2) contributed from clay weathering reactions (such as weathering of K-rich silicate clays such as illite in PASS)(van Breemen, 1973) ; and, (3) ion-exchange reactions.

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The saturation indices for K+ minerals show that the pond pore water is supersaturated with respect to Jarosite-K and thus this mineral will only precipitate (a potential potassium sink).

Given the consistency of K+ concentrations relative to the seawater line and the lack of obvious source points in the pond, it is the authors’ opinion that the K+ is best attributed to run-off containing agricultural fertilisers and subsequent concentration and cycling in the estuary before the water enters the Tomei ponds.

10.7.2 Potassium - spatial distribution in the ponds

In Pond 7, over the three sampling periods, K+ was relatively well mixed throughout the water column (±3%) and across the sediment pore water sample points (Figure 62). In February, the highest concentrations of K+ are in the pore water and water adjacent to the pond dyke walls.

The contoured concentration transects (Figure 62 and Figure 63) show that the range of K+ values in Pond 10 is greater then that in Pond 7. The full range is from 366-442 mg/L which is equivalent to ±10% variability over the entire season. During any of the four sampling periods there is less variability with around ±5% variation across the pond: the ponds are heterogeneous.

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Figure 62: Time series of contoured potassium concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02)

It is worth understanding how the ponds can be heterogeneous with respect to K+ if it is behaving as a conservative element. If K+ concentrations are inherited from the farms intake water and the concentration in the intake water varies with farming practice, rainfall, runoff and tidal circulation: then each time water is taken in, a new concentration gradient is created between the resident pond water and the newly introduced water.

Since the pond mixing has proved inefficient for all other ions it is unrealistic to expect K+ to behave any differently. The measured K+ concentration gradients are evidence that the farms water mixing practices are inefficient. The same

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients heterogeneity is demonstrated by the Cl- concentration transects presented in Figure 69.

Figure 63: Time series of contoured potassium concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)

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10.8 Magnesium

Magnesium can accumulate in the Tomei pond water by: (1) the dissolution of Mg-rich evaporite minerals (carbonate or sulfate); (2) ion-exchange Mg2+ released to solution as Ca2+, 2Na+ or 2K+ are taken up; (3) desorption releasing Mg2+ from clay surfaces in response to environmental chemical change (high cation-exchange capacity clay); (4) the weathering of Mg-rich silicate minerals: olivine, pyroxenes, amphiboles, micas, magnesite, hydromagnesite, brucite and dolomite (Hem, 1989); (5) ocean-derived salty aerosols; or, (6) sea-salts carried in coastal rainwater

Magnesium can be removed from water by: (1) ion-exchange reactions taking up Mg2+ and releasing Ca2+, 2Na+ or 2K+ to the water; (2) adsorption on the surface of clay minerals; and, (3) the precipitation of Mg-rich evaporite minerals during the evaporation of water; or, (4) biological uptake – either plants or animals for nutritional, skeletal or calcareous development.

Magnesium has been found as minor inclusions in the titanium-rich ilmenite grains at Pimpama. They are found in the Carboniferous sandstone deposited as bedrock (Preda and Cox, 2001). This may contribute to small fraction of the Mg2+ dissolved in the pond water.

It is also possible that when dolomite is added to the pond water to neutralise acidic pH (both at the commencement of the season and during the season), Ca2+ is being scavenged by biota for growth, leaving behind an excess dissolved Mg2+.

10.8.1 Observed Magnesium concentrations and potential source/sink

Magnesium versus TDS for Ponds 7 and 10 are plotted in Figure 64. As for potassium, the variations in Mg2+ concentrations fall within the spread of data upon which the seawater line is based. No meaningful geochemical interpretation can be provided to explain differences between concentrations: as this is most likely due to variations in the intake water and a sign of ineffective mechanical mixing (Figure 65).

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Figure 64: TDS versus Mg2+ for (a) Pond 7 and (b) Pond 10

The saturation indices indicate that some Mg-rich minerals are undersaturated and others are supersaturated but no evidence is available to support a conclusion that these reactions are contributing to the variability in observed concentrations.

10.8.2 Magnesium - spatial distribution in the ponds

+ As for K , the concentration heterogeneity is due to ineffective mixing in the ponds (Figure 65 and Figure 66).

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Figure 65: Time series of contoured magnesium concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02)

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Figure 66: Time series of contoured magnesium concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)

10.9 Chloride

Chloride is a conservative element in natural groundwater and surface water systems: once it is dissolved into water, it is not lost from solution by natural geochemical processes. Cl- is removed from water bodies through evaporative

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients concentration (which ultimately leads to Cl-rich mineral precipitation) or its concentration can be reduced by mixing/dilution by lower chloride fluids (e.g. rainwater).

The concentration of chloride is not affected by redox reactions, does not undergo adsorption to mineral surfaces and is not used in biochemical processes. Rainwater close to the coast usually contains several milligrams per litre of Cl-, however this concentration decreases further inland (Hem, 1989): in the context of this study of seawater salinities this rainwater source of Cl- is insignificant.

10.9.1 Data summary and key processes

Chloride is the dominant anion in the seawater and the Tomei dataset. By plotting Cl- against the calculated TDS most of the pond water samples lie on or to the right of the seawater line for both Ponds 7 and 10 (Figure 67). This suggests a slight depletion of Cl- in all months relative to the seawater line. Since Cl- is a hydrochemically conservative ion this can be explained by the precipitation of chloride-rich evaporative minerals. It is important to note that the “most depleted” appear to be in the pore water samples. Chloride-rich mineral precipitation in these sediments is not supported by the saturation indices. Instead this “depletion” is an indicator that non-Cl-ion-bearing chemical reactions in the ponds are controlling pond water chemistry.

Figure 67: TDS versus Cl- for (a) Pond 7 and (b) Pond 10

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Chemical reactions that do not contribute Cl- to the ponds increase the salinity (TDS) of the pond water (Figure 67). These reactions are not active in the estuary and accordingly the seawater line predicts a lower TDS for any given Cl- value. That is, the higher TDS in the pond water samples demonstrates that there are specific reactions proceeding in the ponds and pond sediments that are not occurring in the adjacent estuary.

The saturation indices for Pond 7 suggest that the chloride-rich minerals are all undersaturated or close to being undersaturated (will not precipitate) and this is a likely source of the additional Cl-. Halite is supersaturated in November 2001, but is undersaturated later in the season.

In contrast to this the SI’s for Pond 10 suggests that halite, atacamite and cerargyrite are supersaturated in the month of June, and had the potential to precipitate at the end of the season.

Chloride ion concentrations have fluctuated significantly between the sampling periods (Figure 67), High concentrations were measured during the peak summer months of February and April, and this is due to evaporative concentration of Cl- in both the estuary system and the farms ponds.

10.9.2 Chloride - spatial distribution in the ponds

The concentration transect for Cl- from Pond 7 shows that the pond is heterogeneous with there being a ±5% variation in concentration at any of the sampling periods (Figure 68). Clearly this pond is unaffected by chemical reaction (Cl- being a conservative element) and the heterogeneous nature of the Cl- concentrations further demonstrates the failure of the mechanical pond mixing.

Pond 10 is more homogeneous than Pond 7 in each of the sampling periods (Figure 69).

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Figure 68: Time series of contoured chloride concentration profiles each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02)

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Figure 69: Time series of contoured chloride concentration profiles each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and June-02)

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10.10 Sulfur Species

Sulfur can accumulate in the Tomei pond water by: (1) the dissolution of sulfate-rich evaporite minerals; and more significant in the ASS environment (2) the oxidation of sulfide minerals (e.g. pyrite).

Sulfur can be removed from water by: (1) the precipitation of sulfate-rich evaporite minerals (oxidising conditions); (2) the precipitation of monosulfide or iron sulfide minerals (under strongly reducing conditions)

2- When water is oxidised sulfur is present as SO4 and under acidic conditions this bonds with hydrogen to form H2SO4 (sulfuric acid). When water is reduced sulfur is present in its reduced form of S2-, and depending on the acidity of the 2- - system S progressively transforms with decreasing pH into HS and H2S.

In poorly mixed/aerated, ASS affected aquaculture ponds as the season extends DO will drop as ASS reacts and the residual oxygen is consumed by pond biota, in parallel H2S will rise and progressively the ponds aquaculture species will experience increased stress which in turn will result in a deterioration of pond productivity.

10.10.1 Data summary and key processes

2- 2- In Ponds 7 and 10 S is common in the pore water samples and SO4 is common in the water column samples (Figure 70). Sulfate concentrations are lower in Pond 10 than in Pond 7 and this is consistent with Pond 7 being more 2- ASS affected. In both cases SO4 is present in concentrations 1000-2000 times that of S2-. It is sensible that S2- is higher in the pore waters as there are 2- 2- more reduced: if oxygen were available S will transform rapidly into SO4 .

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2- 2- Figure 70: S versus SO4 (meq/L) for (a) Pond 7 and (b) Pond 10

Pond 7 has water column samples containing the reduced form of sulfur (S2- see Figure 71) whereas Pond 10 does not: reflecting the better efficiency of the mechanical aeration in Pond 10.

Figure 71: TDS versus S2- (mg/L) for (a) Pond 7 and (b) Pond 10

2- Figure 72 shows that SO4 (the oxidised form of sulfur) is present in both ponds at concentrations that are consistent with the seawater line. A small 2- number of samples from Pond 7 have excess SO4 in solution. These are derived from the pore water and water column samples. Pond 10 has fewer 2- high concentration SO4 samples than Pond 7. This difference is attributed to reactions occurring in the more ASS-rich sediments of Pond 7, and the subsequent migration of those ASS leachates into the overlying water column.

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2- Figure 72: TDS versus SO4 for (a) Pond 7 and (b) Pond 10

10.10.2 Sulfur - spatial distribution in Pond 7

Figure 73 presents the concentration transect for Pond 7. During the first 2- sampling round in November SO4 concentrations were relatively low. In 2- February the sand pore water in the centre of the pond had a very high SO4 concentration (peak of 7,141 mg/L), with the overlying water column having 2- relatively low concentrations. In April, at the end of the season, SO4 is still concentrated in the centre of the pond, but has migrated vertically from the sediments into the overlying water column.

2- It is interesting that there is a SO4 peak (Figure 73) in the sediments in the centre of the pond in February and in the adjacent sediments (closer to the pond walls) there are S2- peaks (Figure 74). This is consistent with the bivariate plots presented in Figure 72 showing that the February sampling round identified peak sulfur concentrations, and that this was the only month 2- where SO4 was consistently above the seawater line.

This peak in sulfur coincides with the hottest/driest month and is most likely linked to warmer water and sediments allowing accelerated reaction rates. The 2- oxidised form (SO4 ) being concentrated in the centre of the pond and the reduced form (S2-) being on its flanks is not readily explained. There are two possible mechanisms: (1) oxygenated groundwater influx/up welling in the centre of the pond providing oxygen to drive sulfide oxidation, or (2) the Tomei ponds have a subtle mound in the centre of the pond and if this had a higher elevation than the adjacent sampling points the chemical change may be

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients highlighting stratification of the pond oxygen base with the pond centre being above and the flanks below this horizontal plane.

2- Figure 73: Time series of contoured SO4 concentration profile for each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02)

Figure 74 presents the time series of S2- concentrations in Pond 7. In all three sampling rounds peak S2- concentrations are located in the pond sediment. As the season progresses the S2- concentrations progressively increase. As the season progresses the sediments become more anoxic as the ponds biological demand for DO increases, and as a result the reduced species dominates. However there is a simultaneous ongoing inorganic oxidation reaction (pyrite oxidation) as the ASS reacts with the relatively oxygenated pond water. This reaction produces acidity and consumes any available dissolved oxygen, as a result the released sulfur takes on the form of S2- and is

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients rapidly transformed into H2S(g). Complimentary evidence of the progress of this reaction was provided by the strong H2S odour observed during the collection of the sediment pore water samples.

Figure 74: Time series of contoured S2- concentration profile for each sampling round as measured in Pond 7 (Nov-01, Feb-02 and Apr-02)

10.10.3 Sulfur - spatial distribution in Pond 10

2- 2- Figure 75 and Figure 76 present the concentration transects for SO4 and S 2- in Pond 10. A quick comparison with the SO4 transects from Pond 7 shows that pond 10 is more homogenous and appears to have better connectivity between the sediments and the water column.

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2- Figure 75: Time series of contoured SO4 concentration profile for each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and Jun-02)

2- As observed in Pond 7 the highest concentration of SO4 was measured in 2- February and was from the sand pore water (with a SO4 concentration of 3,649 mg/L). In Pond 10 we do not see the same build-up and migration of a

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2- SO4 hot spot from the sediments into the ponds water column as we do in Pond 7: this is largely put down to (1) their being better hydraulic connectivity between the sediments and the water column and (2) the Pond 10 mechanical mixing and aeration devices being more effective.

It is clear from Figure 76 that as the season progresses there is a build-up of S2-concentrations in the pond sediments. November has very little S2-, then in February there is a peak in the middle of the pond, in April this migrates to the flanks of the pond, and in June this is almost evenly distributed across the pond.

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Figure 76: Time series of contoured S2- concentration profile for each sampling round as measured in Pond 10 (Nov-01, Feb-02, Apr-02 and Jun-02)

Figure 77 clearly demonstrates the relationship between sample location (water column or pore water) and pH. In Pond 7, all of the water column samples have a pH above 7.7; and the pore water samples have a pH below 7.7. In Pond 10 the water column samples have a pH above 8.2, and all the pore waters have a lower pH. Likewise all elevated S2- concentrations are in the pore water samples (Figure 78).

2- Figure 77: SO4 (mg/L) versus pH for (a) Pond 7 and (b) Pond 10

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Figure 78: S2- (mg/L) versus pH for (a) Pond 7 and (b) Pond 10

It is interesting to note that Pond 7 has a larger amount of acidic pH’s than Pond 10 by approximately 0.5 pH units: Pond 7 is more influenced by the underlying ASS.

Figure 79 shows that S2- in Pond 7 is only present when DO is below 3.5 mg/L; in Pond 10 the cut off is 2.0 mg/L. It is also worthy of note that S2- was only present in the water column in Pond 7.

Figure 79: S2- (mg/L) versus DO for (a) Pond 7 and (b) Pond 10

10.11 The Carbonate System

The author recognises that there are both organic and inorganic carbon sources in water systems. This thesis only considers the inorganic carbon system as the project funding limited the available scientific flexibility.

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Dissolved Inorganic Carbon (DIC) can accumulate in water by: (1) the dissolution of carbonate-rich minerals; (2) biological activity (respiration - plants and organisms use oxygen and release CO2 into the water); (3) atmospheric exchange (air and rainfall).

Carbon can be removed from water by: (1) the precipitation of carbonate-rich minerals; (2) biological activity (photosynthesis - plants remove CO2 from the water in the presence of sunlight and chlorophyll and produce oxygen); (3) atmospheric exchange (out gassing)

The form that the DIC takes is dependent on the pH of the environment (Figure

80) – under acidic conditions (pH<6.4) DIC is predominantly CO2 and H2CO3; - at mid-range pH (6.410.33) CO3 is dominant.

Figure 80: Bjerrium plot (Drever, 1997)

Carbon should be dealt with as one group. This is best done by calculating DIC and Total Alkalinity (Alk) using hydrochemical modelling software, and then this can be plotted against the relevant chemical parameters.

Total Alkalinity is expressed as presented in Equation 19 and Dissolved Inorganic Carbon is expressed as shown in Equation 20: the difference being the inclusion of carbonic acid and the stoichometry of carbonate.

- 2- Equation 19 Total Alkalinity = m HCO3 + m 2CO3 - - 2- Equation 20 Dissolved Inorganic Carbon (5CO2) = m H2CO3 + m HCO3 + m CO3

Where m = the molar concentration of the species (Drever, 1997; Neal, 2001).

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In an open system DIC=Alk as the partial pressure of CO2 (PCO2) is controlled by the atmospheric CO2 concentration.

10.11.1 Dissolved Inorganic Carbon versus Total Alkalinity

The comparison of DIC to Alk is predominantly done to establish whether the system is open or closed with respect to CO2(g). By defining the stoichometry of the key reactions that influence the generation of CO2(g) Deffeyes (1965) was able provides a plot of these variables to assist interpretation.

When plotted using DIC (Figure 81 a-f) there is little difference between Pond 7 and Pond 10. The most notable difference is that Pond 7 has a DIC peak that is almost twice that of Pond 10.

Figure 81a and Figure 81d show that DIC is independent of the chloride concentration of the water and is therefore not simply accumulating in the pond waters. It is clear from the same graphs that the elevated DIC values relate to the sand and ASS pore water samples (are associated with the pond sediments).

Figure 81b and Figure 81e show that DIC and pH are closely related: the - highest DIC’s are at pH’s of between 7 and 8 (where HCO3 is the dominant carbon species). All of the elevated DIC samples are from the sediment (sand and ASS) pore waters.

Figure 81c and Figure 81f are the standard Deffeyes (1965) plot which shows that there is a difference in the DIC source for Pond 10. Pond 7 data plots predominantly on the 1:1 line, whereas in Pond 10 the data is more clearly spread across the 1:1 and sulfate reduction line, with a subset of the June samples plotting toward the methanogenesis line.

The 1:1 line represents equilibrium between calculated Total Alkalinity and calculated Dissolved Inorganic Carbon: this is the expected outcome for a pond open to the atmosphere with good atmospheric CO2(g) exchange. The data plotting on the sulfate reduction line shows that there is CO2(g) being generated and released into the pond by the action of sulfur reducing bacteria (Equation 21). The deviation toward the methanogenesis line indicates that

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Chapter 10: Pond Hydrochemistry - Major Ions & Nutrients there is no oxygen availability and all potential sulfate reduction is complete, and the methanogenic bacteria have commenced metabolising organic matter and nutrients to form CO2(g) and CH4(g) (see Equation 22).

Figure 81: Chloride, pH and Alkalinity verses DIC for Pond 7 (a, b, c) and Pond 10 (d, e, f) 2- + + 2- Equation 21 (CH2O)106(NH3)16H3PO4 + 53SO4 + 67H = 106CO2 + 16NH4 + HPO4 +106H2O + + 2- Equation 22 (CH2O)106(NH3)16H3PO4 + 14H = 53CO2 + NH4 + 53CH4 + HPO4

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The strongly reduced environment shown by the June data from Pond 10 is a direct reflection of the post harvest timing of sampling (after the aerators/agitators have been turned off and the pond is being prepared for draining). Clearly these mechanical agitators and aerators have been effective in providing sufficient oxygen throughout the season to keep the pond in equilibrium with the atmosphere (DIC: Alk at 1:1), however as shown by the pond transects above they have not delivered optimal mixing (chemical homogenisation) or sufficient oxygen to adequately support the farms aquaculture species.

10.11.2 Carbon Speciation

- At the pH of the pond waters HCO3 is the dominant carbon species. Inspection of the horizontal scale on Figure 83 is one order of magnitude greater than that 2- of CO2 (Figure 82) and two orders greater than that of CO3 (Figure 84).

Pond 7 has CO2 in solution in the pond water but there is no recorded CO2 in the water column of Pond 10 (Figure 82).

Figure 82: CO2 (mg/L) versus depth for (a) Pond 7 and (b) Pond 10 (note: the red dashed line represents the sediment-water interface)

- Pond 7 has on average, slightly more (251.01 mg/L) HCO3 in solution in the pond water than is present in Pond 10 (203.10 mg/L) (Figure 83).

2- Neither Pond 7 nor Pond 10 (Figure 84) have appreciable CO3 in the 2- sediment pore waters. Pond 10 has approximately double the CO3 concentration in the water column.

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- Figure 83: HCO3 (mg/L) versus depth for (a) Pond 7 and (b) Pond 10 (note: the red dashed line represents the sediment-water interface)

2- Figure 84: CO3 (mg/L) versus depth for (a) Pond 7 and (b) Pond 10 (note: the red dashed line represents the sediment-water interface)

The proportional changes in carbonate species described above are a direct function of the pH of the ponds: Pond 10 is more alkaline than Pond 7. This geochemical observation fits well with the geological observation that Pond 7 is underlain by a higher proportion of ASS/PASS.

10.12 Nutrients

Nutrients in the ponds are dominated by nitrogen and phosphorus. Nitrogen - + was measured as the redox species NO3 (Nitrate) and NH4 (Ammonium: + measured as NH3-N and recalculated as NH4 ). Phosphorus was measured as 3- phosphate (PO4 ).

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Nutrient concentrations are intentionally elevated by farm management to encourage the production and maintenance of phytoplankton blooms. These nutrients are generally delivered to the pond water as agricultural fertilisers. Agricultural fertilisers are generally made up of variable ratios of Nitrogen (N), Phosphorus (P) and Potassium (K). Uneaten prawn feed, excrement and algae also contribute to the nutrient load in the prawn ponds: however this builds up through accumulation and so is more progressive than the instantaneous nutrient spikes associated with the addition of fertiliser.

The following two reactions are dominant in determining the species and concentration of Nitrogen in the pond water:

Nitrification as seen in Equation 23 (Freeze and Cherry, 1979):

+ - + Equation 23 O2 + 1/2NH4 = 1/2NO3 + H + 1/2H2O

Denitrification as seen in Equation 24 (Freeze and Cherry, 1979):

- - + Equation 24 CH2O + 4/5NO3  2/5 N2(g) + HCO3 + 1/5H + 2/5H2O

Most of the P in the pond water is also derived from the fertilisers; however there is also a contribution from the prawn feed; from the effluent discharging from the adjacent sugar cane farms; or, from the dissolution of deposited P minerals. Temporetti and Pedrozo (2000) estimated that about 66% of P in fish feed accumulates on the bottom sediments. The P associated with these sediments can be released to the water column when pH drops and promotes P-rich mineral dissolution.

10.12.1 Pond 7

A summary of the range and average concentrations for the nutrients from Pond 7 are presented in Table 20.

Table 20: Descriptive statistics for Nutrients in Pond 7 (in mg/L) (n=105)

+ - 3- NH4 NO3 PO4 Pond 7 (mg/L) (mg/L) (mg/L) Water Column Average 10.22 10.31 0.38 Pore water 22.62 16.59 1.81 Water Column Min 0.58 1.32 0.00 Pore water 7.80 4.40 0.04 Water Column Max 47.09 25.52 4.30 Pore water 88.69 86.24 5.60

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The analytical data and a corresponding interpretation is provided in the sections below.

10.12.1.1 Ammonium

+ - Ammonium is the reduced species of the NH4 /NO3 redox couple. Nitrification + - occurs when NH4 is oxidised to form Nitrate (NO3 ) according to the reaction defined in Equation 25:

+ - + Equation 25 O2 + ½ NH4  ½ NO3 + H + ½ H2O

From this balanced reaction it is clear that the nitrification generates H+ which will acidify the host solution.

Ammonium concentrations are higher in the pore water than in the water column (Figure 85a) and in Pond 7 this is associated with pH values less than + 7.7. The highest NH4 concentrations in the dataset are in the most saline samples (Figure 85b) where salinities are high largely due to peak summer heat and solar radiation causing evaporative concentration in the ponds and adjacent estuary.

+ + Figure 85: (a) pH versus NH4 (b) TDS versus NH4 for Pond 7

+ Figure 86a and Figure 86b show that NH4 concentrations are highest in the most reduced pore water samples (when the Eh< -300 and the DO<2 mg/L).

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+ + Figure 86: (a) Eh versus NH4 and (b) DO versus NH4 for Pond 7

10.12.1.2 Nitrate

Like Ammonium, Nitrate is present in higher concentrations in the pore waters than the water column (Figure 87a) and is reaches its highest concentrations in the most saline samples (Figure 87b).

- - Figure 87: (a) pH versus NO3 (b) TDS versus NO3 for Pond 7

- Figure 88a and Figure 88b show that the concentration of NO3 is greater when the Eh is less than -250mV and the DO is less than 2.0 mg/L.

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- - Figure 88: (a) Eh versus NO3 (b) DO versus NO3 for Pond 7

- The Pond 7 samples have elevated concentrations of NO3 when compared to the seawater line. This is most likely the result of the farmer’s practice of adding fertiliser to the ponds to promote algal blooms. It is also possible that there is a contribution from the estuary where fertilisers applied to the cane crops has runoff as wastewater. Once the water enters the estuary, the flanking mangroves will help to regulate the amount of nitrogen in the estuary.

10.12.1.3 Phosphorus

3- The greatest concentrations of PO4 occur in the pore waters when the pH is 3- below 7.7 (Figure 89a). There higher concentration of PO4 in the sediment 3- pore water suggests that PO4 is concentrating in the pond floor sediment. This inference is supported by the work of Briggs and Funge-Smith (1994) who concluded that the sediments are the primary sink of P in the majority of aquaculture ponds.

3- In Figure 89b and Figure 90 the concentration of PO4 is greater when the Eh is less than -100mV and the DO Is less than 3.0 mg/L.

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3- 3- Figure 89: (a) pH versus PO4 (b) Eh versus PO4 for Pond 7

3- Figure 90: DO versus PO4 for Pond 7

The data from Pond 7 suggest that the nutrient load for both nitrogen and phosphorus are the result of the farmer’s addition of fertilisers to the pond in an attempt to promote algal blooms.

10.12.2 Pond 10

A summary of the range and average concentrations for the nutrients from pond 10 are presented in Table 21.

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Table 21: Descriptive statistics for Nutrients in Pond 10 (in mg/L) (n=134) NH + NO - PO 3- Pond 10 4 3 4 (mg/L) (mg/L) (mg/L) Water Column Average 9.26 10.68 0.14 Pore water 13.83 31.74 1.62 Water Column Min 4.06 0.01 0.00 Pore water 5.16 0.00 0.00 Water Column Max 14.90 14.08 15.09 Pore water 42.25 156.64 178.35

The analytical data and a corresponding interpretation is provided in the sections below.

10.12.2.1 Ammonium

As in Pond 7, the Pond 10 pore water samples that have the highest ammonium concentrations, are the most acidic, most reduced and most saline samples in the dataset (Figure 91a, Figure 91b, Figure 92a, Figure 92b).

+ + Figure 91: (a) pH versus NH4 (b) TDS versus NH4 for Pond 10

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+ + Figure 92: (a) Eh versus NH4 (b) DO versus NH4 for Pond 10

+ The NH4 concentrations in both Pond 7 and 10 plot on (or close to) the seawater line (Figure 85b and Figure 91b) on this basis Tomei is unlikely to be contributing any significant Ammonium load to the adjacent estuary system.

10.12.2.2 Nitrate

+ As in Pond 7, and as for NH4 , the Pond 10 pore water samples have the highest nitrate concentrations, are the most acidic, most reduced and most saline samples in the dataset (Figure 93, Figure 94a and Figure 94b).

- Figure 93 shows that the data from November and February have higher NO3 + concentrations than the seawater line. During the same months NH4 was on the seawater line so it is unlikely that this enrichment is the result of nitrification as the stoichiometery of Equation 25 consumes and produces a molar ratio of 1:1.

It is likely that (1) given that these are the hottest months with the most solar - radiation and (2) the elevated concentration is as NO3 (oxidised form of nitrogen as is available in powered fertilisers) the chemical data is pointing to the increase in concentration being due to the farmer adding fertiliser to promote algal blooms to protect the pond prawns from heat stress.

The phytoplankton would use sunlight and the fertiliser to photosynthesise and this would support/boost the pond DO. This elevated DO was noted in the pond during November and February (Figure 94a) sampling rounds. The - presence of this DO ensured that NO3 remained available in solution rather + than being denitrified back to NH4 .

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Figure 93: TDS versus NO3 for Pond 10

Figure 94: (a) DO versus NO3 (b) Eh versus NO3 for Pond 10

- Figure 94a and Figure 94b show that the concentration of NO3 is greater when the DO is less than 2 mg/L and the Eh is less than -200mV.

10.12.2.3 Phosphorus

The interpretation of this dataset is made difficult by the presence of the two 3- 3- high PO4 outliers. Figure 95a shows that there is generally increasing PO4 with increasing (more alkaline) pH.

3- This may be explained by PO4 concentrations responding to the dissolution of phosphate-rich minerals such as vivianite (Fe3(PO4)2.8H2O) (see Figure 95b) as it is undersaturated in all water samples (water column and pore waters).

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3- Figure 95: (a) pH versus PO4 (b) Vivianite Calculated Saturation Indices for Pond 10

3- 3- Figure 96: (a) Eh versus PO4 (b) DO versus PO4 for Pond 10

Figure 96a and Figure 96b suggest that there is a general trend for increased 3- PO4 concentrations in oxidising conditions (the June pond water samples).

As summated for Pond 7 the nutrient concentrations in Pond 10 are largely controlled by the farm practice of fertilising the ponds to stimulate algal blooms. This farm practice impacts the pond chemical system by resulting in an - increase in DO which in this case has supported an elevated NO3 concentration in the pond water.

10.13 Chemical Characteristics of Water - Statistical Analysis

10.13.1 Spearman Correlation Coefficients

The Spearman correlation coefficient (which measures how closely variables are related) for Ponds 7 and 10 are referred to in Appendix 7.

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10.14 Summary

The major cations indicate that Ponds 7 and 10 are chemically similar with the following basic behaviour:

• Excess sodium is probably derived from 2Na-Ca ion exchange, and Na- rich carbonate/sulfate mineral dissolution.

• Calcium concentrations are controlled by Ca-Carbonate dissolution, ion exchange and uptake by biota.

• Potassium and Magnesium are acting conservatively in the ponds.

It is clear that the mechanical mixing that was employed at the time of this study was ineffective in homogenising the ponds.

The major anions also indicate that Ponds 7 and 10 are chemically similar with both ponds the following basic behaviour:

• Chloride as a proportion of the TDS of the samples is present in concentrations slightly less that is predicted from the seawater line. This is evidence that there are active geochemical reactions proceeding in the study ponds which contribute to the chemistry and TDS (salinity) of the ponds.

• Sulfur is present in Ponds 7 and 10 in concentrations that are similar to 2- 2- that of the seawater line. S is common in the pore waters and SO4 is common in the water column. The oxidation of sulfide minerals is a key geochemical process in these ponds. Sulfur concentrations are lower in Pond 10 than in Pond 7 which is consistent with Pond 7 being more ASS/PASS affected.

• Carbonate is present in the ponds in concentrations consistent with or slightly higher than seawater. In the acidic pore waters carbonate dissolves, in the pond water it precipitates or is taken up by biota, Carbonate speciation is pH controlled with the pH of pond 10 being higher (more alkaline) than that of pond 7.

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The ponds have a higher salinity (up to 42ppt) than “normal” seawater (35ppt); this is most significant in the summer months due to the effect of evaporative concentration in the estuary (impacting intake waters) and the ponds.

In both ponds pore water samples have the highest nutrient concentrations, are the most acidic, most reduced and most saline samples in the dataset. These nutrients are believed to be largely derived from the farming practice of adding fertiliser to the ponds to promote the growth of algae. The relative contribution of nutrients from the anthropologically nutrient enriched estuary system (through the intake waters) can not be meaningfully quantified with the available dataset.

Multivariate statistics (Spearman Correlation Coefficients) were run on the fully hydrochemical dataset and these methods readily identified the dominant elements common in seawater: Na+Cl and Sr+B.

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11 POND HYDROCHEMISTRY - TRACE ELEMENTS

11.1 Introduction

Trace elements associated with ASS are largely derived from the weathering of clays and the dissolution of the minerals that make up the host sediments and rocks. Acid Sulfate Soils generate acid when weathering and this accelerates the weathering, dissolution and mobilisation of trace metals. Once mobilised, trace elements can be stable in solution, can complex with other ions and/or precipitate from solution as secondary minerals.

In this chapter trace elements are discussed in the following order: abundant trace elements (Al, Mn, Fe), transition metals (V, Cr, Co, Ni, Mo), alkaline earths (Sr, Ba), other transition metals (Cu, Ag, Zn), metalloids (As, B) and radioactive elements (U).

Forty seven trace elements were analysed for this study using ICP-MS. The high salinity of the samples led to a requirement to dilute each sample at a ratio of 1:30 with Milli-Q water prior to analysis (Jarvis, 1992). This dilution has resulted in the concentration of some elements in the diluted samples being below the ICP-MS instruments detection limits. The concentrations of some elements in the analysed water samples exceed the ANZECC (2000) Toxicant Guidelines for the Protection of Aquaculture Species for Saltwater Production. Table 22 provides a summary of the guidelines and the dilution corrected detection limit (for each analysed trace element) of the ICP-MS instrument used in this study.

Table 22: Combined table of ANZECC Guidelines (2000) for maximum concentrations of elements and detection limits of the ICP-MS used in this study Measured parameter Symbol/ INORGANIC DL** Notation TOXICANTS* ()g/L) Aluminium Al <10 0.21 Ammonia (un-ionised) NH3 <100 - Arsenic As <30 0.04 Barium Ba - 0.04 Boron B - 0.66 Cadmium (varies with hardness) Cd <0.5–5 - Chlorine Cl <3 - Chromium Cr <20 0.09 Cobalt Co - 0.01 Copper (varies with hardness) Cu <5 0.16 Hydrogen sulfide H2S <2 - Iron Fe <10 -

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Lead (varies with hardness) Pb <1–7 - Magnesium Mg - - Manganese Mn <10 0.02 Molybdenum Mo - 0.10 Nickel Ni <100 0.06 - - Nitrate (NO3 ) NO3 <100,000 - Nitrite (NO2) NO2 <100 - Phosphates PTOT <50 - Selenium Se <10 - Silver Ag <3 0.03 Strontium Sr - 0.01 Total Available Nitrogen (TAN) TAN <1,000 - Uranium U - 0.00 Vanadium V <100 0.15 Zinc Zn <5 0.23 * ANZECC Saltwater Production Guideline ()g/L) ** Dilution adjusted Detection Limit ()g/L) of ICP-MS instrument used in this study

11.2 Abundant Trace Elements

11.2.1 Aluminium

Hem (1989) states that Al is the third most abundant element in the earth’s crust, however it is usually only dissolved in natural water in minute concentrations unless the system is acidic (pH<4).

Aluminium is typically attributed to the dissolution of silicate minerals such as feldspars, feldspathoids, mica and amphiboles. When silicates weather they release major cations into solution and form Al-rich clays (e.g. kaolinite). Subsequent chemical weathering of these clays (particularly under acidic conditions – a common feature of ASS) can result in mobilisation of Al into pond pore and water column water.

The literature demonstrates that there is a tendency for aluminium in estuarine sediments to associate with dissolved organic matter (Pardue and Patrick 1995; Mackin and Aller, 1984).

11.2.1.1 Concentrations and Distribution in the ponds

The concentrations of AlTOT in both Pond 7 and 10 exceed the ANZECC Guidelines (by more than 5 times in Pond 7, and 3.5 times in Pond 10) across most of the dataset.

In Pond 7 the water column samples have higher AlTOT concentrations than those of the sediments (Figure 97a). In Pond 10 there is a nearly equal

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Chapter 11: Pond Hydrochemistry - Trace Elements concentration of AlTOT in the water column and the pond sediment pore waters (Figure 97b).

In Pond 7 the highest recorded concentration of AlTOT was 1,044.33μgl (at a pH of 7.91); this is clearly an outlier with the next highest sample being 189.63μgl (at a pH of 8.07). In Pond 10 the highest recorded concentration of AlTOT was 233.85μgl (at a pH of 8.54). Only the outlier in Pond 7 is higher than the concentrations in the Pimpama estuary and farm intake waters (pink crosses in Figure 97a and b).

Figure 97: (a) AlTOT versus TDS for Pond 7 (note: the red dotted line represents the ANZECC (2000) trigger value for Al TOT). All data to the right of the red line exceeds the guideline. The two pink seawater crosses to the right of the red dotted line are samples taken from the Tomei intake water and estuary. The two pink seawater crosses to the left of the line are reference seawater from the GBR and Hem. (b) AlTOT versus TDS for Pond 10

Hem (1989) stated that AlTOT is only concentrated in acidic aqueous environments (pH<4). pH and AlTOT are cross plotted in Figure 98a and b and this clearly shows that AlTOT is in solution over the whole pH range in the Tomei ponds (6.5-9.0). Similarly the elevated AlTOT in the Pimpama estuary waters are at a pH of 8.0-8.3.

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Figure 98: (a) pH versus AlTOT for Pond 7, (b) pH versus AlTOT for Pond 10

11.2.1.2 Potential source minerals

Excluding the one outlier in Pond 7, the Pimpama estuary and intake drain waters are carrying concentrations of AlTOT in solution that are 3-5 times greater than that present in either Pond 7 or 10. Since the Pimpama estuary is the source of the water for water exchanges in the Tomei ponds there must be an effective AlTOT sink in the Tomei ponds.

Bigham and Norsdstrom (2000) found that Al-hydroxysulfate minerals precipitate when acidic water is rapidly neutralised. This process may assist by mitigating the influx of Al into the pond when Al-rich ASS leachates enter the water column from the pond base and walls. On this basis minerals such as basaluminite (Al4(OH) 10SO4), jurbanite (Al(SO4)(OH)) and alunite

(KAl3(SO4)2(OH)6) should go from undersaturated in the pore waters to supersaturated in the water column. This is not the case in Ponds 7 and 10: all three minerals are supersaturated in the pore water samples and undersaturated in the water column (see Figure 99 and Figure 100).

In Pond 7 the low ALTOT in the pore waters in November is likely to be due to the precipitation of Al(OH)3(a), basaluminite (Al4(OH) 10SO4), jurbanite

(Al(SO4)(OH)), and to a lesser extent alunite (KAl3(SO4)2(OH)6) (Figure 99). The sink for AlTOT in the pond water as well as the sediments (when compared to the intake waters) is likely to be due predominantly to the precipitation of diaspore (AlOOH), and to a lesser extent boehmite (AlOOH), and gibbsite

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(Al(OH)3) as these two phases are undersaturated in the water column during February

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Figure 99: Saturation Indices for Aluminium bearing minerals in Pond 7

In Pond 10 (Figure 100) the June pore water samples are supersaturated with respect to each of the presented mineral phases, as a result any combination of these phases may provide an AlTOT sink in the pore waters at that time. Focussing only on the other sampling periods: diaspore (AlOOH), gibbsite

(Al(OH)3) and to a lesser extent boehmite (AlOOH) are supersaturated in the pore waters. Alunite (KAl3(SO4)2(OH)6), is near saturation and may either dissolve or precipitate in the sediments. Diaspore (AlOOH) is the only mineral phase that is at or above saturation in the water column and is the most likely ALTOT sink in pond 10.

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Figure 100: Saturation indices for Aluminium bearing minerals in Pond 10

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11.2.2 Manganese

Manganese can substitute for iron, magnesium, or calcium in silicate structures. It has three valence states Mn2+, Mn3+ and Mn4+: the water soluble state is Mn2+ (Hem 1989; Spokes and Liss, 1995).

Manganese is often derived from rocks: (1) it is a constituent of many igneous and metamorphic rock forming minerals such as olivine, pyroxene and amphibole; (2) it has been identified in titanium-rich ilmenite grains from Pimpama’s Carboniferous sandstone bedrock (Preda and Cox, 2001); and (3) it can be derived from the weathering of dolomite and limestone (Hem, 1989).

In the presence of oxygen and elevated pH Mn will precipitate as a Mn4+ surface crust. These encrustations can contain co-precipitated iron, cobalt, lead, zinc, copper, nickel and barium. Manganese will also selectively precipitate on an existing manganese oxide surface.

Manganic oxides are rapidly reduced in sediments (Equation 26 and Equation 27) and are a common source of manganese ions in reduced pore waters (Jorgensen, 1983). Under acidic conditions (like those created by sulfide oxidation) Manganese can accumulate in water to create high concentrations (Golez and Kyuma, 1997). Krom and Sholkovitz (1978) analysed anoxic pore waters in sediments from Loch Duich, Scotland. They found that dissolved Mn concentrations are the highest at the sediment-water interface, and peak concentrations were associated with the anoxic-oxic redox boundary.

+ 2+ - Equation 26 CH2O + 2MnO2(s) + 3H = 2Mn + HCO3 + 2H2O (Reduction) 2+ Equation 27 O2 + Mn + 2H2O = 2MnO2(s) + 4H2O (Oxidation)

Manganese is biologically important as it is integral in photosynthesis, in detoxifying oxygen products during respiration, and used in the production of ATP. However, manganese can only be used by biota in its reduced form (Spokes and Liss, 1995).

11.2.2.1 Concentrations and Distribution in the ponds

Figure 101a and Figure 101b show that the MnTOT concentrations in both ponds significantly exceed the ANZECC Guidelines (2000). Figure 101a and b also show that concentrations of MnTOT in solution in Ponds 7 and 10 are

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Chapter 11: Pond Hydrochemistry - Trace Elements highest in the pore water samples. In Figure 102a and b, MnTOT is in solution when pH is between 6.5 and 8.0.

Figure 101: MnTOT concentrations versus TDS for (a) Pond 7 (Note: the red dotted line shows the ANZECC Guidelines (2000) trigger concentration) and (b) Pond 10

Figure 102: pH versus MnTOT for (a) Pond 7 and (b) Pond 10

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Figure 103: DO versus MnTOT for (a) Pond 7 and (b) Pond 10

Figure 104: Eh versus MnTOT for (a) Pond 7 and (b) Pond 10

Manganese concentrations are highest when the available oxygen (whether measured as Eh or DO) is low (see Figure 103 and Figure 104) clearly demonstrating the solubility/mobility of Mn under reduced and acidic (Figure 102) conditions.

The Spearman correlation coefficient shows that MnTOT does not have a strong positive relationship with any other trace metals. The closest relationship is with Fe2+ and Fe3+ (0.564 and 0.554 respectively) (Appendix 7 and Figure 105a and Figure 105b).

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Figure 105: (a) MnTOT versus Fe2+ for Pond 7 and (b) MnTOT versus Fe3+ Pond 7

For Pond 10, the Spearman correlation coefficients (Appendix 7) show again (as in Pond 7) that MnTOT is strongly correlated to reduced Fe (0.715), but unlike in Pond 7, it also correlates with Ba (0.884).

11.2.2.2 Potential source minerals

Figure 101 shows that the Pond 7 and 10 pore water samples contain higher concentrations of MnTOT than either the Pimpama Estuary or intake samples. There is clearly a source for Mn in the ponds. It is likely that this source for MnTOT is the mobilisation of Mn (dissolution of Mn-rich minerals in reduced acidic conditions that are created as a by-product of pyrite oxidation (ASS oxidation)) from the pond base and dyke walls.

The Mn concentrations in the ponds water column are similar to that of the Pimpama estuary and intake waters and this is because in the presence of (1) elevated pH (saltwater buffering) and (2) oxygen (atmospheric contact) the Mn is not stable in solution, and forms a precipitate that settles on the pond floor. This natural mechanism mitigates Mn concentrations in the pond water column and ensures that the Tomei farm does not discharge Mn-rich water into the estuary system, but does not assist the sediment dwelling P. japonicus prawn.

Manganese is highly mobile and therefore can be transported into the pond from acidifying pond walls and reduced pond bottom sediments. Once it enters the pond water, it is readily precipitated or adsorbed because of its instability under oxic, neutral to alkaline pH conditions.

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The Pond 7 SI’s (Figure 106) show that the minerals that are able to precipitate from the water column in November and April were hausmannite

(MnO4), Mn3AsO4.8H2O, bixbyite (Mn2O3), and in November only pyrochroite

(Mn(OH)2). In February Rhodochrosite (MnCO3) was approximately at saturation during all sampling rounds and had the potential to continually precipitate any excess Mn from the water column.

Rhodochrosite (MnCO3) and MnHPO4 are supersaturated in the pore water samples however there is unlikely to be sufficient CO3 available at the sediment pH (carbon being present as CO2 or HCO3 see Figure 80) however there is ample PO4 (Section 10.12.1.3 and 10.12.2.3) available and adequate acidity for MnHPO4 to precipitate.

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Figure 106: Saturation Indices for manganese minerals in Pond 7

In Pond 10 (Figure 107) the SI data indicate that all Mn-bearing minerals (other than Rhodochrosite (MnCO3)) are undersaturated in all months except June.

MnHPO4 is supersaturated in the pore water samples and is likely to be in the sediments. It is likely that Rhodochrosite precipitation controls water column Mn concentrations.

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Figure 107: Saturation indices for Manganese minerals in Pond 10

11.2.3 Iron

The weathering of the minerals that constitute igneous rocks (pyroxenes, amphiboles, biotite, magnetite, and nesosilicate olivine) produce ferrous (Fe2+) iron and to a lesser extent Ferric (Fe3+) iron (Hem, 1989). ASS and their pore

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Chapter 11: Pond Hydrochemistry - Trace Elements waters are Fe-rich containing iron sulfide minerals where fresh; and iron oxides where they have been weathered.

Iron in the ponds at Tomei was measured in the field as the reduced species Fe2+ (using a spectrophotometer) and later in the laboratory Total Fe was measured by ICP-OES.

Fe3+ was determined by calculating the difference between FeTOT and the Fe2+ measurements (Luther, 1995), however the recorded total iron was inconsistent with the reported values often being lower than the measured Fe2+ concentrations from the field. This is put down to iron-oxide precipitation in the sample bottle (during sample storage) prior to the FeTOT analysis.

- 2- - Iron (lll) usually complexes with ligands such as OH , HPO4 and F , forming colloid and particulate oxyhydroxides. Iron (III) is mobile in acidic sediment pore water as a ferric-organic (humic-fluvic) complex (

In reduced (anoxic) conditions Fe2+ is the dominant form of iron and in the presence of sulfide can precipitate as Fe-sulfide. Fe2+ is thermodynamically unstable and is soluble in water, however when exposed to oxygen it oxidises rapidly. Like Fe3+, Fe2+ is able to complex with organic matter when the acidity of the system is appropriate (Krom and Sholkovitz, 1978; Spokes and Liss, 1995; Rose and Waite, 2003).

11.2.3.1 Concentrations and Distribution in the ponds

The greatest concentration of Fe2+, Fe3+ and FeTOT in both Pond 7 and 10 are located in the pore water from the pond sediments (Table 23).

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Table 23: Descriptive statistics for Fe2+, Fe3+ and total iron (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Pond 7 Pond 10 Statistic Fe2+ Fe3+ FeTOT Fe2+ Fe3+ FeTOT (mg/L) (mg/L) (mg/L) (mg/L) (mg/L) (mg/L) Water Column Average 0.02 0.00 0.02 0.02 0.00 0.02 Pore water 0.83 0.52 1.35 0.93 1.02 1.95 Water Column Min 0.00 0.00 0.00 0.00 0.00 0.00 Pore water 0.00 0.00 0.00 0.01 0.00 0.01 Water Column Max 0.17 0.07 0.21 0.40 0.00 0.40 Pore water 5.10 3.99 9.09 7.10 13.30 20.40

Total Iron in both ponds increases with decreasing pH (Figure 108). Figure 109 shows that there is not a direct relationship between increasing TDS and Fe2+ concentration.

Figure 108: pH versus FeTOT in (a) Pond 7 and (b) Pond 10 (use legend on (a) for both graphs)

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Figure 109: TDS versus Fe2+ for (a) Pond 7 and (b) Pond 10 Table 24: Iron concentrations from the seawater samples taken near Tomei (note: see Figure 2 for reference seawater sample location map)

Sample Location pH FeTOT (mg/L) 20m from farm intake/discharge point in channel 7.98 0.42 50m from junction of farm channel and estuary 8.18 <0.05 Junction between estuary and ocean at Jumpinpin Bar 8.30 <0.05

Figure 109 (and comparison of Table 24 and Table 23) show that there is a larger amount of Fe2+ in the Tomei ponds pore waters than in the Pimpama drain, estuary or seawater samples. This indicates that there is a source of iron in the pond sediments and this source is reacting with the pore waters.

The Spearman correlation coefficients (Appendix 7) show that the ions that are most closely correlated to the Fe species are the other Fe species and to a lesser extent MnTOT (Pond 7 = 0.550 and Pond 10 = 0.715). In Pond 10, FeTOT also positively correlates with Ba (0.676).

11.2.3.2 Potential source/sink minerals

High concentrations of Fe in these pond sediments is best explained in the ASS environment by the dissolution of pyrite and other Fe-rich minerals that are present in the sediments. The oxidation of pyrite consumes available oxygen, produces Fe2+, sulfate and H+ ions (generates acidity: see Equation 2).

Figure 110 shows the relevant Fe-bearing mineral SI’s for Pond 7. The pore waters are supersaturated in all months. Chalcopyrite and pyrite are supersaturated in both the water column and pore water samples across the entire grow-out season. These two minerals were identified by XRD as being a component of the pond sediment samples (see 7.4.3).

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Figure 110: Saturation Indices for iron minerals in Pond 7

In Pond 10 pyrite and chalcopryrite are again super saturated in the water column and the sediments (Figure 111). Mackinawite and FeS (ppt) are undersaturated in the water column and supersaturated in the sediments with a collection of samples being on the saturation line.

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Figure 111: Saturation indices for iron minerals in Pond 10

It is important to note that the SI plot for pyrite in both Pond 7 and 10 shows that a small number of the sediment pore waters plot on (or very close to) the saturation line: at this line the samples have the thermodynamic potential to either oxidise or precipitate pyrite. As the water column is strongly supersaturated, it is thermodynamically feasible that pyrite is precipitating from the water column and is being deposited on the pond floor. Intuitively this is unlikely as the DO of the water column is too high for the precipitation of pyrite. However, the activity of this (or a similar) Fe-scavenging reaction is evidenced by the relatively low concentrations of total iron in solution.

In the oxidised portions of the ponds Fe3+ is the predominant iron species. Fe

(III) is associated with oxides and hydroxides (e.g. Fe(OH)3). In the more reduced portions of the pond the dominant iron species is Fe (II). As pore

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Chapter 11: Pond Hydrochemistry - Trace Elements waters become depleted of oxygen (anaerobic) and become more acidic ferric hydroxide minerals begin to dissolve according to Equation 28:

+ 2+ Equation 28 ½ Fe2O3 (s) + 2H (aq) + ¼ CH2O(s)  Fe (aq) + 5/4 H2O(aq) + ¼ CO2(g)

Note that for each four moles of Fe (II) produced there is a mole of CO2(g).

Figure 112 shows that CO2 concentrations are elevated in the sediment pore waters. The dissolution of undersaturated iron hydroxides is the potential origin of this CO2.

Figure 112: Depth versus CO2 for (a) Pond 7 and (b) Pond 10

11.2.3.3 Impact on the surrounding estuary

Inspection of Figure 109 and Table 24 shows that the Tomei intake water channel has a Fe2+ concentration marginally higher than that of either Pond 7 or 10. This is reasonable evidence to support the hypothesis that Fe concentration control is occurring in the ponds and that Tomei farm is not releasing Fe rich waters into the estuary system.

The drain sample is different to either the estuary or the Jumpinpin bar samples as neither of these contain detectable Fe. It is likely that the iron has been lost from solution as it has entered the estuary system either through (1) oxidation and precipitation as discussed above, or (2) biological uptake.

Blue green algae are known to scavenging iron during both photosynthesis and nitrogen fixation (Paerl and Tucker, 1995; Emmenegger et al., 1998). The complexing of soluble iron by oxidation and subsequent precipitation is

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Chapter 11: Pond Hydrochemistry - Trace Elements decreased in the presence of organic ligands. This allows Fe to remain in the dissolved phase for a greater time and (once algae break the organic ligands) to be bioavailable for uptake by cyanobacteria. There have been documented outbreaks of cyanobacteria (blue green algae) such as Lyngbya sp. (majuscula) (also known colloquially as Mermaids Hair or Fire Weed) in Moreton Bay (Preda and Cox, 2002). This alga contains stinging cells that, if a person comes into contact with, can cause irritation of the skin and eyes. If the algae is accidentally ingested or inhaled (during swimming) it can cause irritation to the digestion or respiratory tract (www.epa.qld.gov.au/environmental_management/coast_and_oceans/marine_ha bitats/lyngbya_updates/facts_and_contacts, 2004). The alga affects aquatic fauna (and possibly mobile flora) as it is also toxic to fish, to the point where they will avoid areas where the Lyngbya grows.

11.2.3.4 Impact on the aquaculture species

Penaeus japonicus prawns bury themselves in the sediment to evade predators. The sediment of Pond 7 is low in dissolved oxygen (negative Eh), has a relatively low pH and is therefore likely to be uncomfortable for prawns. This is however a suitable environment for the accumulation of Fe2+ (Equation 2 and the relevant sections of Chapter 9).

During osmoregulation, the prawns inadvertently transform the Fe2+ (reduced) into Fe3+ (oxidised) on the surface of their gills. As a result Fe oxyhydroxides accumulate on their gills; this reduces the surface area for oxygen absorption, and limits prawn respiration, ultimately leading to suffocation of the prawns (Govinnage, 2001).

11.3 Transition Metals

11.3.1 Vanadium

Vanadium is found in plants, coal and petroleum and is involved in the biochemical processes of living organisms (Hem, 1989). Vanadium (V) has three oxidation states (V3+, V4+, and V5+) with the dominant form being V5+. In this study Vanadium was measured as total V.

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Mandiwana et al. (2005) suggested that vanadium associated with the sediment fraction, can be transferred to water during the weathering of shale and clay minerals. The concentration of V leached from sediment is directly 3- proportional to the concentration of PO4 in the sediment (Mandiwana et al., 2005) and this can mean that V concentrations can increase with the application of phosphate-rich fertilisers.

Vanadium can be removed from water through adsorption to iron oxides in sediment (Poledniok and Buhl, 2003), or can co precipitate with iron, illite or smectite, but has limited mobility unless in acidic conditions (Preda and Cox, 2001).

11.3.1.1 Concentrations and Distribution in the ponds

Table 25 and Figure 113 show that there is no discernable difference between pore water and water column V concentrations. It is however interesting that on the whole V in the ponds is lower than in the Pimpama estuary or intake drain (Appendix 4 and 5).

Table 25: Descriptive statistics for vanadium (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Statistic Pond 7 Pond 10

V V (μg/L) (μg/L) Water Column Average 67.64 55.30 Pore water 70.15 56.51 Water Column Min 33.87 26.22 Pore water 24.03 35.61 Water Column Max 156.66 86.58 Pore water 221.13 78.21

The ANZECC Guidelines (2000) trigger value for saltwater aquaculture is <100 μg/L however, only 7.6% of the Pond 7 data exceeds this concentration and 70% of those are sampled from the water column (Figure 113a). There is less V in solution in Pond 10 (average = 55.79) than in Pond 7 (average = 68.74): none of the Pond 10 samples exceed the ANZECC Guidelines (2000) for Vanadium (Table 25 and Figure 113b).

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Figure 113: Total vanadium versus TDS (with ANZECC Guidelines trigger value marked on in a red dotted line) (a) for Pond 7, and (b) for Pond 10

11.3.1.2 Potential source/sink minerals

Iron-rich ilmenite from the Carboniferous sandstone bedrock of the Pimpama region is known to contain Vanadium (Preda and Cox, 2001).

Most of the samples from both ponds contain Vanadium at concentrations lower than that of the farm intake waters suggesting that there is a Vanadium sink in the ponds. Eh abundance of iron in the system (related to oxidation of ASS) suggests it is likely that the V sink is a combination of both adsorption to ion oxides and or scavenging driven by the precipitation of iron rich minerals. The pH of Pond 10 is higher (more alkaline) than that of Pond 7 making V less mobile in pond 10 and leading to the lower measured V concentrations in Pond 10.

Based on the Spearman correlation coefficients for Pond 7, Vanadium has a strong positive correlation with As (0.791) (Figure 114a), along with strong relationships with B (0.852), Cr (0.767), and Li (0.776) (Figure 115). In Pond 10 the Spearman correlation coefficients show the same relationship with As (0.805) (Figure 114b) as well as B (0.669), Li (0.577) and a further correlation between V and both Rb (0.777), and Sr (0.629).

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Figure 114: Vanadium versus arsenic in (a) Pond 7 and (b) Pond 10

Figure 115: Vanadium compared to (a) Boron, (b) Chromium and (c) Lithium in Pond 7

11.3.2 Chromium

Chromium is typically associated with ultramafic igneous rock forming minerals or their weathered counterparts (such as chromite, FeCr2O4, which is highly

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Chapter 11: Pond Hydrochemistry - Trace Elements resistant to weathering). In rock forming minerals, chromium’s oxidation state is usually Cr3+ (low mobility and non-toxic), however Cr6+ is the water soluble form, it is highly mobile and toxic (mutagen and carcinogen in mammals). Chromium was measured as total Chromium and no attempt was made at individually analysing the Cr species.

Chromium can be associated with organic ligands, smectite and can be oxidised in the presence of Fe and Mn(IV) oxides and hydroxides and its mobility is pH dependent with pH 4.5-6.0 being the optimal range (Hem, 1989; Shanker et al., 2005; Astrom, 2001; Preda and Cox, 2001; Stepniewska et al., 2004).

Preda and Cox (2001) suggests that in the Pimpama region Cr may be present in oxides associated with the basement rock units (though ionic substitution for Fe and Al). They also suggested that there is Cr in the sediments (greywacke, shale and chert) of the Neranleigh-Fernvale Beds. They measured 28-43ppm of Cr in the greywacke, 26-52ppm in shale and 5-13ppm in the chert.

11.3.2.1 Concentrations and Distribution in the ponds

There no discernable difference between the average concentration of CrTOT in the water column and pore water samples from the same months (Figure 116 and Table 26).

Table 26: Descriptive statistics for Chromium in Pond 7 (Pond 7 n=105)

Sample Location Descriptive Statistic Total Chromium (μg/L) Water Column Average 8.26 Pore water 9.31 Water Column Minimum 5.28 Pore water 6.09 Water Column Maximum 15.63 Pore water 18.15

The ANZECC Guidelines (2000) trigger value for saltwater aquaculture is 20 μg/L and samples from Pond 7 did not exceed the guidelines (Figure 116).

The Pimpama estuary and intake waters plot in the centre of the Chromium data clusters for the warmer months (November and February) (Appendix 4 and 5), however the cooler months of April and June show relative enrichment

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Chapter 11: Pond Hydrochemistry - Trace Elements of Cr. In Pond 7 the April data is split in half, a cloud of data near the concentration of the intake waters and the other half showing close to 100% enrichment (Figure 116a). In Pond 10 all of the April data is enriched relative to the intake waters by approximately 100%, in contrast the June data is split into a lower concentration cloud and a higher concentration cloud (with approximately 40% enrichment: see Figure 116b).

Figure 116: Total chromium versus TDS (with ANZECC Guidelines trigger value marked on in a red dotted line) (a) Pond 7 and (b) Pond 10

11.3.2.2 Potential source minerals

There are greater concentrations of CrTOT in the pond water and the Pimpama seawater samples than in the GBR and Hem seawater samples. This suggests that Cr is being contributed to the Pimpama estuary from the areas geological units (basement and Neranleigh-Fernvale Beds described above).

The relative enrichment of CrTOT in some of the April and June samples is more difficult to explain. Review of the DO data provide an insight – the enriched samples correlate with higher DO and a positive measured Eh (Figure 117a and b). This correlation suggests that there is possible Cr liberation from an unidentified ligand, perhaps in response to the precipitation of Mn and Fe oxides.

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Figure 117: Eh versus Total chromium (with pink dotted line to show the transition from anoxic to oxic waters) (a) Pond 7 and (b) Pond 10

In Pond 7 Chromium strongly correlates with V (0.767) and has a weaker correlation with As (0.791), B (0.578) and Li (0.487) (Appendix 7).

11.3.3 Cobalt

Cobalt is found as Co2+ or Co3+, it tends to act like Fe and can substitute for Fe in ferromagnesian rock minerals (Hem, 1989). Cobalt has been found in iron- rich ilmenite inclusions in Pimpama’s Carboniferous sandstone bedrock. Cobalt is highly mobile in acidic conditions. Preda and Cox, (2001) identified cobalt in the Pimpama River after rainfall and suggested that it was derived from the weathering of ASS on the river banks.

11.3.3.1 Concentrations and Distribution in the ponds

The ANZECC Guidelines (2000) for saltwater aquaculture do not provide a trigger value for Co.

There is no notable difference between the concentrations of Co in water column and pore water in either pond (Figure 118). The Pond 7 samples either contain no detectable Co or have concentrations that are consistent with the intake waters. There is a Co sink in the system which is actively scavenging Co to reduce pond concentrations in both the pore waters and water column.

In Pond 10 the November samples have some Co concentrations similar to that of the Pimpama intake water and others that are below detection: again

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Chapter 11: Pond Hydrochemistry - Trace Elements showing the effectiveness of a Co sink (particularly in February). Later in the season (April and June) Co concentrations have increased four fold (Figure 118b): at this time there is a significant Co source in the ponds.

Figure 118: TDS versus Total Cobalt for (a) Pond 7 and (b) Pond 10

Cross plotting of Co against the waters physical variables suggest that decreasing Co concentration correlates with higher measured Eh (Figure 119). Note that the blue dashed line on both graphs has the same scale and describes the Co depletion pathway in the low concentration samples for both ponds. The green line on Figure 119b suggests that the oxidation is also controlling Co concentrations in the high Co samples. Figure 120 shows that Co and pH share a similar relationship to that of Co and Eh, with this graph emphasising the difference between the more acidic pore waters and the more alkaline water column.

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Figure 119: Eh versus Total Cobalt for (a) Pond 7 and (b) Pond 10

Figure 120: pH versus Total Cobalt for (a) Pond 7 and (b) Pond 10

The strongest positive Spearman correlation coefficients in Pond 7 are between Co and Sr (0.524). The strongest Spearman correlation coefficients in Pond 10 are between Co and Sr (0.682) with a weaker correlation between Co and Ag (0.545) (Appendix 7). Cobalt may be associated with Sr as a function of the depositional environment of the Neranleigh-Fernvale shale: a deep-sea fan (Lohe, 1980). Figure 121 shows that the Spearman correlation coefficient is picking up the Co-Sr relationship in the low Co concentration samples from both ponds.

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Figure 121: Strontium versus Total Cobalt for (a) Pond 7 and (b) Pond 10

11.3.3.2 Potential source minerals

The Co in the pond water at Tomei is most likely from the weathering of ASS in the ponds. Inspection of Figure 119 and Figure 120 shows that Co is highest in the Pond 10 pore waters: this points to the source being the pond sediments. Pond 7 is known to be more ASS affected, and again the highest Co concentration was identified in the sediment pore waters: it is however worth noting that this is inconsistent with the overall Co concentrations in solution being lower in Pond 7.

11.3.4 Nickel

2+ Nickel has a common form of Ni and also forms oxides (NiO2) and hydroxide

(Ni(OH)3) complexes. Like cobalt, Ni can also substitute for Fe in ferromagnesian igneous rock minerals and silicate dominated soil structures, and can co precipitate with Fe and Mn oxyhydroxide sediments (Rose and Ghazi, 1998; Preda and Cox, 2001). Nickel with manganese are known to co precipitate in the Pacific Ocean and Hem (1989) suggests that this mechanism regulates seawater Ni concentrations.

11.3.4.1 Concentrations and Distribution in the ponds

The ANZECC Guidelines (2000) trigger value for saltwater aquaculture is <100 μg/L: all of the samples from Pond 7 and all but one from Pond 10 are below threshold.

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There is no discernable difference between the Ni concentrations in the intake waters or the ponds, nor is there a discernable difference between the pore water and water column concentrations (Figure 122). There are a small number of pore water samples from pond 7 and 10 which are enriched in nickel and these are generally related to the November sampling round. Figure 123 shows that there is no consistent link between the pH of these pore waters and the elevated Ni concentrations.

Figure 122: Total Nickel versus TDS for (a) Pond 7 and (b) Pond 10

Figure 123: pH versus Ni for (a) Pond 7 and (b) Pond 10

In Pond 7, nickel has a strong positive correlation with Cu (0.744) and mild positive correlations with Al (0.583), As (0.581), Mo (0.623) and Zn (0.568) (Appendix 7). The Spearman correlation coefficients show that Ni is strongly positively correlated with Cu (0.744) and mild positive correlations with Mo (0.631) and Zn (0.642).

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11.3.4.2 Potential source minerals

Nickel in the estuary system is derived from mineral weathering (measured Ni range from 6-27ppm) of the geological units across the Pimpama region (Preda and Cox, 2001).

Nickel in the Tomei ponds is generally behaving conservatively with only a few of the pore water samples showing that in pond reactions are contributing Ni to solution.

11.3.5 Molybdenum

Molybdenum is found in the oxidation states of Mo3+, Mo4+, Mo6+. Above pH 5, 2- the dominant form of Mo is MoO4 and the sulfide ore Molybdenite (MoS2). Mo is a key component of the nitrogen cycle (it is used by bacteria that promote nitrogen fixation). Preda and Cox (2001) found that Mo was not concentrated in the water samples from ASS. They suggested that this may be due to Mo’s low mobility under acidic conditions.

11.3.5.1 Concentrations and Distribution in the ponds

The ANZECC Guidelines for saltwater aquaculture do not provide a trigger value for Molybdenum. Molybdenum does not show a significant difference in concentration between the water column and the pore water samples. Table 27 presents the descriptive statistics for this element for Pond 7.

Table 27: Descriptive Statistics for Molybdenum in Pond 7 (n=105)

Sample Location Descriptive Statistic Total Molybdenum (μg/L) Water Column Average 4.90 Pore water 3.16 Water Column Minimum 0.00 Pore water 0.00 Water Column Maximum 13.95 Pore water 11.34

In Pond 10 the average concentration of Mo is greater than in Pond 7. The pore water maximum Mo concentration is 3 times higher in Pond 10 than in Pond 7 (Figure 124).

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The Spearman correlation coefficients show that there is a positive correlation between Mo and Cs (0.672), Cu (0.631), Ni (0.623), U (0.546) in Pond 7. In Pond 10 this is between Mo and Ni (0.582); and Mo and U (0.511). Preda and Cox (2001) noted a similar relationship between Mo and Ni in their samples and suggested that it related to the weathering of hornblende containing traces of Mo, Co, Ni, Cu and Zn.

11.3.5.2 Potential source minerals

The highest Mo concentrations are those samples from the Pimpama drain, and estuary and the pond 10 pore waters at the commencement of the growout season (Figure 124). The intake waters contain a higher concentration of Mo than the other seawater reference samples (GBR and Hem). This suggests that there is a source of Mo contributing to the elevated concentration in the Pimpama estuary system.

Figure 124: Mo versus pH for (a) Pond 7 and (b) Pond 10

The most likely source of Mo in the Pimpama estuary system may be due to: (1) weathering from the catchments geology or (2) run off from fields that have had fertiliser applied (3) contamination from heavy industry.

The slight depletion of Mo in the ponds water column (relative to the intake waters) may be due to the uptake of nitrogen fixing bacteria (see Section 10.12).

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11.4 Alkaline Earth Metals

11.4.1 Strontium

The average seawater concentration of Sr is 8 mg/L (Hem, 1989). The source of Sr in the pond water is most likely related to the seawater; however, Sr can also replace Ca or K in the mineral lattice of silicate minerals (which comprise igneous rocks).

11.4.1.1 Concentrations and Distribution in the ponds

A trigger value for Sr is not listed in the ANZECC Guidelines for saltwater aquaculture as this element is naturally found in seawater in concentrations of about 8000 μg/L (Hem, 1989).

There is little difference between the average concentrations for the water column and pore water (Figure 125 and Table 28).

Table 28: Descriptive Statistics for Strontium (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Statistic Pond 7 Pond 10 Total Strontium (μg/L) Total Strontium (μg/L) Water Column Average 3,951.94 4,149.74 Pore water 3,656.46 4,352.28 Water Column Minimum 2,100.66 1,973.88 Pore water 1,970.49 2,000.91 Water Column Maximum 10,953.80 10,075.30 Pore water 7,149.06 9,530.34

Pond 7 and 10 are generally depleted with respect to Sr compared to the seawater reference samples (Figure 125).

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Figure 125: Chloride versus Sr (mg/L) for (a) Pond 7 and (b) Pond 10

There is a positive correlation between Sr and As (0.715), B (0.760), Co (0.524), Li (0.697), Rb (0.722) and V (0.578) in Pond 7 (Appendix 7). The Spearman correlation coefficients for Pond 10 show positive correlation between Sr and Ag (0.530), As (0.705), B (0.599), Rb (0.761) and V (0.629).

11.4.1.2 Potential source minerals

The Pimpama seawater samples are similar to the GBR and Hem concentrations showing that Sr (in the Pimpama estuary system) is behaving as a conservative element (Figure 125). This is not the case in the ponds: there are a limited number of samples which show Sr enrichment (relative to the intake water) in Pond 7 in February and Pond 10 in November and June; with the rest of the samples showing Sr depletion.

Figure 126: pH versus Sr for (a) Pond 7 and (b) Pond 10

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Figure 127: DO versus Sr for (a) Pond 7 and (b) Pond 10

2- Figure 128: CO3 versus Sr for (a) Pond 7 and (b) Pond 10

2- Figure 129: SO4 versus Sr for (a) Pond 7 and (b) Pond 10

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From Figure 126 there is not a clear relationship between pH and Sr concentration. Figure 127 suggests that there is a tendency for less Sr in solution with increasing DO. Figure 128 shows that Sr concentrations generally decrease with increasing Carbonate in solution. Figure 129 shows that Sr concentrations decrease with increasing Sulfate in solution.

It is likely that Sr is being lost from solution in the ponds through the precipitation of either SrCO3 (Strontianite) and/or SrSO4 (Celestite).

11.4.2 Barium

Barium is associated with igneous and sedimentary rocks (shale, sandstone and carbonate). In sediment, naturally occurring Ba is generally contained in the minerals barite (BaSO4) and barium carbonate (BaCO3).

Barium carbonate (witherite: BaCO3) has a similar solubility to calcite (Hem,

1989). The largely insoluble barium sulfate (barite: BaSO4) is considerably more soluble in the presence of chloride when pH is less than 9.3 (http://www.atsdr.cdc.gov/toxprofiles/tp24-c5.pdf).

Barium is strongly adsorbed to clay minerals and can substitute in the mineral lattice with K+. Barium can be removed from the water through adsorption on the surfaces of metal oxides or hydroxides.

Insecticides can contain barium fluorosilicate and carbonate (http://www.atsdr.cdc.gov/toxprofiles/tp24-c5.pdf) and this is a potential anthropological contribution of Ba to the environment.

Conditions such as pH, Eh, CEC, and the presence of sulfate, carbonate, and metal oxides affect the partitioning of Ba and its compounds in the environment. Generally, the solubility of barium compounds increases with decreasing pH (http://www.atsdr.cdc.gov/toxprofiles/tp24-c5.pdf).

11.4.2.1 Concentrations and Distribution in the ponds

A trigger value for Ba is not included in the ANZECC Guidelines for saltwater aquaculture; however this element is typically present in seawater with a concentration of approximately 20 μg/L (Hem, 1989).

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Table 29 presents the statistics for Ba in the ponds. This table shows that in pond 7 there is a two-fold, and in Pond 10 a four-fold difference between the maximum concentrations of Ba in the water column and the pore water.

Table 29: Descriptive Statistics for Barium (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Statistic Pond 7 Pond 10 Total Barium (μg/L) Total Barium (μg/L) Water Column Average 20.46 21.12 Pore water 68.91 87.92 Water Column Minimum 7.92 6.00 Pore water 7.05 22.35 Water Column Maximum 120.57 111.78 Pore water 214.05 439.59

Figure 130 shows that there is a higher concentration of Ba in the pore waters than in the water column and that Ba concentrations are highest in both ponds when pH is low. In both ponds there is one outlier sample with a Ba concentration that is approximately twice that of all other samples (Figure 130).

Figure 130: pH versus Total Ba in (a) Pond 7 and (b) Pond 10

In Pond 10 the Ba concentrations are similar to those of Pond 7 but are approximately 20 μg/L higher for any given pH.

There is a positive correlation between Ba and Y (0.608) in Pond 7 with weaker correlations between Ba and Fe (0.464) (Appendix 7). In Pond 10 Ba is positively correlated to As (0.629), reduced Fe (0.676), Total Fe (0.558), Mn

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(0.884) and Y (0.537); and weakly correlated to U (0.441), V (0.450) and Zn (0.426).

11.4.2.2 Potential source minerals

The hydrochemical data shows that the concentration of Ba in the Pimpama intake drain is higher (Ba is over 90 μg/L) than that in adjacent estuary (12 μg/L) or reference seawater (~20 μg/L): there is a Ba source feeding the Pimpama intake sample point that is not affecting the adjacent estuary. Figure 130 shows that the intake sample is approximately the same as the highest pond water column Ba concentrations.

This intake channel provides both the discharge channel for Tomei and the adjoining sugar cane farm. The similarity between the maximum water column concentrations and the intake water sample suggests a large portion of the excess Ba in solution is derived from the Tomei ponds.

The elevated Ba in the pond sediment pore waters demonstrates that there is a reaction occurring in the pond sediments which is releasing Ba. These sediments are relatively acidic (ASS) and are strongly reduced. It is likely that: - (1) witherite (BaCO3) is dissolving to produce Ba and HCO3 /CO2(g) (species depends on acidity) in response to the in-situ ASS generated acidity; and, (2) 2- - barite (BaSO4) is undergoing reduction to release Ba and S /HS /H2S (species also depends on acidity).

11.5 Other metallic elements

11.5.1 Copper

Copper is associated with the greywacke, shale and chert units that make up the bedrock of the Pimpama region (Preda and Cox, 2001). Copper can replace Fe and Mg in hornblende and mica, and K, Na or Ca in feldspars. Copper is a highly mobile element in acidic aqueous environments.

11.5.1.1 Concentrations and Distribution in the ponds

The ANZECC Guidelines (2000) trigger value for Cu in saltwater aquaculture is <5 μg/L. The Pimpama estuary and drainage channel samples as well as most

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Chapter 11: Pond Hydrochemistry - Trace Elements of the samples from Pond 7 and Pond 10 exceed these guidelines (Figure 131). Clearly the Pimpama estuary is exposed to a significant Cu source.

In both ponds there is little difference between the water column and pore water Cu concentrations (Table 30).

Table 30: Descriptive Statistics for Copper (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Statistic Pond 7 Pond 10 Total Copper (μg/L) Total Copper (μg/L) Water Column Average 8.51 6.90 Pore water 7.26 6.33 Water Column Minimum 0.00 0.00 Pore water 0.00 0.00 Water Column Maximum 79.56 56.79 Pore water 17.40 27.69

The majority of samples in Ponds 7 and 10 contain smaller concentrations of Cu than the Pimpama estuary and intake samples (Figure 131): there is a Cu sink in the ponds.

Figure 131: Total Cu versus TDS in (a) Pond 7 and (b) Pond 10

In Pond 7 there are positive Spearman correlation coefficients between Cu and Al (0.599), As (0.577), Cs (0.526), Mo (0.631), Ni (0.744), V (0.571) and Zn (0.642) with weaker correlations with B (0.483) and Li (0.411) (Appendix 7).

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11.5.1.2 Potential source minerals

Figure 132 shows that at the ponds pH there is little variation in Cu concentration suggesting that concentration is not controlled by pH.

Figure 132: Total Cu versus pH in (a) Pond 7 and (b) Pond 10

There is however a clear suppression of the intake Cu concentrations in the ponds. The Cu sink is most likely chalcopyrite which the SI data indicate is thermodynamically prone to precipitation from the pond water column and pore waters (see Section 11.2.3).

11.5.2 Silver

The most common oxidation state of silver is +1 (silver nitrate: AgNO3); less common is the +2 form (silver (II) fluoride: AgF2) or +3 compounds (silver (III) persulfate: Ag2(SO5)3). Silver is found in native form, combined as ores with sulfur (argentite: Ag2S), arsenic, antimony, or chlorine (cerargyrite/horn silver: AgCl).

Silver in the Pimpama estuary is likely transported as mineralised inclusions in rock fragments washed downstream from mineral deposits around Mount Tambourine.

11.5.2.1 Concentrations and Distribution in the ponds

The descriptive statistics for AgTOT in ponds 7 and 10 are presented in Table 31. Figure 133 shows that many samples exceed the ANZECC Guidelines (<3 μg/L), in both ponds at least one sample forms an outlier having Ag

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Chapter 11: Pond Hydrochemistry - Trace Elements concentrations 20 to 30 times higher than recommended by the guidelines (Figure 133).

Table 31: Descriptive Statistics of AgTOT (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Statistic Pond 7 Pond 10 Total Silver (μg/L) Total Silver (μg/L) Water Column Average 2.85 2.20 Pore water 0.79 6.70 Water Column Minimum 0.00 0.00 Pore water 0.00 0.00 Water Column Maximum 64.77 15.09 Pore water 4.32 178.35

Figure 133: AgTOT versus TDS (with ANZECC Guidelines marked on) for (a) Pond 7 and (b) Pond 10

Figure 134 shows that the highest Ag concentrations are recorded when pH is around 8.0.

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Figure 134: AgTOT versus pH for (a) Pond 7 and (b) Pond 10

In Pond 7 AgTOT has a positive relationship with most trace elements, with U being the closest positively related element (0.511). Manganese, Fe3+ and FeTOT have a weak negative relationships and Fe2+ has no linear relationship (0.00) with AgTOT. In Pond 10, AgTOT is positively correlated with Sr (0.530) (Appendix 7). This may suggest that in Pond 10, Ag concentrations are associated with reactions involving a Sr-bearing mineral.

11.5.2.2 Potential source minerals

The AgTOT in the Pimpama estuary and drain waters are 0.1 μg/L and 0.3 μg/L respectively: significantly lower than the pond water column and most pore water concentrations. There is a source of AgTOT in the pond.

It is possible that the silver source is (1) a naturally occurring silver-bearing mineral which is dissolving under the pond conditions; or, (2) there is an anthropogenic silver source (e.g. introduced to the farm inadvertently by farm practices – trucked in sediments, chemical usage, or feed).

The saturation indices for cerargyrite (AgCl) are strongly undersaturated and this mineral will (if present in the pond sediments) dissolve (source) to contribute silver. Similarly Ag metal ranges from undersaturation to supersaturation so may dissolve or precipitate. Saturation Indices for Ponds 7 and 10 indicate that Acanthite (AgS) is supersaturated (Figure 135 and Figure 136) in all samples from both ponds and is a potential silver sink.

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Figure 135: Saturation Indices for Acanthite in Pond 7

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Figure 136: Saturation Indices for silver minerals in Pond 10

11.5.3 Zinc

Zinc is a relatively common element and is consumed by plants and animals. It is generally in the form of Zn2+.

Preda and Cox (2001) found that Zn at Pimpama is commonly associated with greywacke, shale and chert. Zinc can also substitute for Fe/Mn in silicates.

Zinc concentrations are influenced by ion adsorption, ion-exchange or co precipitation (Hem, 1989). Preda and Cox (2002) stated Zn is a dominant trace metal in the Pimpama system and that it shows weak correlations with minor metals as a result of co precipitate with Fe or Mn oxides, and clays (smectite and illite). Zinc is mobile under oxidising and acidic conditions, but it can be immobilised as organic complexes in sulfides or in sediments (Preda and Cox, 2002).

Anthropogenic sources of Zn are generally related to the liming of ASS soils, the corrosion of metal structures (from acidic runoff associated with ASS) and from heavy industry.

11.5.3.1 Concentrations and Distribution in the ponds

The ANZECC Guidelines (2000) trigger value for Zn in <5 mg/L. Inspection of Table 32 and Figure 137 shows that there is no discernable difference in pore water or water column Zn concentrations, and that most samples from both ponds have Zn concentrations greater than the trigger level.

Table 32: Descriptive Statistics for Zinc (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Statistic Pond 7 Pond 10 Total Zinc (μg/L) Total Zinc (μg/L) Water Column Average 21.09 18.89 Pore water 19.24 12.60 Water Column Minimum 0.00 0.00 Pore water 0.00 0.00 Water Column Maximum 104.46 307.86 Pore water 47.79 42.24

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Figure 137: ZnTOT versus TDS (with ANZECC Guidelines in red) (a) Pond 7 and (b) Pond 10

The Pimpama estuary water and intake drain waters have Zn concentrations of 22 and 26 μg/L respectively. This is slightly higher than the average concentrations in the pond and is approximately in the middle of the data clusters for each pond. The range of this cluster is significant (0 - 60 μg/L) indicating that there are both zinc sources and sinks in both ponds.

The Spearman correlation coefficient data for Pond 7 and 10 shows that Zn does not have a strong relationship with any other trace metal. There is a weak positive correlation with Cu (0.642) and Ni (0.568) in Pond 7 and the following ions in Pond 10: Ba (0.426), Mo (0.464), Ni (0.538), and U (0.562) (Appendix 7).

11.5.3.2 Potential source/sink minerals

Minerals such as sphalerite (ZnS), wurtzite (ZnS) and Zn5(OH)8Cl2 are likely to be actively cycling and co precipitating in the pond as conditions change on a daily basis in response to changes in pond geochemistry: providing both a source and a sink for Zn with isolated pockets of the pond accumulating high Zn concentrations while others are relatively depleted.

11.6 Metalloids

Metalloids are elements classified as having the properties of both metals and non-metals (Dictionary of Science, 1971).

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11.6.1 Arsenic

Arsenic can exist in two primary species in sediments and water, they are: arsenate (As(V)) and arsenite (As(lll)).

Arsenic solubility is greater under anaerobic conditions (and is in the form of As (lll)). Arsenic (V) dominates oxygenated environments (Pardue and Patrick, 1995). Arsenic samples in this study are sampled as AsTOT.

11.6.1.1 Concentrations and Distribution in the ponds

The ANZECC Guidelines (2000) trigger value for AsTOT is <30 μg/L. The mean AsTOT concentrations for both the water column and pore water exceed the by 2 and the greatest concentration of AsTOT measured in the pond is over 4 times the acceptable concentration (Figure 138).

The mean concentrations of AsTOT in the water column and the pore water of both ponds are very similar (Table 33).

Table 33: Descriptive Statistics for Arsenic in Pond 7 (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Statistic Pond 7 Pond 10 Total Arsenic (μg/L) Total Arsenic (μg/L) Water Column Average 57.38 53.29 Pore water 60.13 59.56 Water Column Minimum 28.80 21.48 Pore water 21.42 40.98 Water Column Maximum 109.86 97.17 Pore water 130.20 84.45

There are lower maximum concentrations AsTOT in Pond 10 than in 7. However, most of the data collected from Pond 10 exceed the ANZECC Guidelines (2000).

Figure 138 suggests that there is a decrease in As with increasing salinity in both Ponds 7 and 10. In both Pond 7 and 10 the AsTOT is lower than that of the Pimpama estuary or intake channel samples. There is an arsenic sink in these ponds.

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Figure 138: AsTOT versus TDS for (a) Pond 7 and (b) Pond 10

In Figure 139 AsTOT concentration does not vary across a range of pH values showing that these are largely independent variables.

Figure 139: pH versus AsTOT for (a) Pond 7 and (b) Pond 10

Spearman correlation coefficients (Appendix 7) for Pond 7 highlights that As is positively correlated to B (0.912), Li (0.837), Rb (0.832) and V (0.791) (see Figure 114). In Pond 10 As is positively correlated to B (0.770), Ba (0.629), Li (0.670), Rb (0.818), Sr (0.705), U (0.624) and V (0.805).

11.6.1.2 Potential source minerals

All of the pond water data has a lower AsTOT concentration than the Pimpama estuary samples. AsTOT is being lost from the pond water: the sink is likely to be the precipitation of As-bearing mineral(s) and/or adsorption reactions involving Fe oxyhydroxides.

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11.6.2 Boron

Boron is again derived from igneous rocks and can also be found in biotite and amphibole minerals. Seawater has about 4.6 mg/L of B and this acts as a chemical buffer in the oceans (Hem, 1989).

11.6.2.1 Concentrations and Distribution in the ponds

The ANZECC Guidelines (2000) do not provide a trigger for B. The average concentration of B in both ponds in the water column and the pore water are very similar (Table 34).

Figure 140 shows that the ponds have a wide range in B concentrations when compared to the Hem (1989) average seawater concentration. The Pimpama estuary and intake waters however have concentrations close to Hem’s (1989) average.

Table 34: Descriptive Statistics for Boron (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Statistic Pond 7 Pond 10 Total Boron (μg/L) Total Boron (μg/L) Water Column Average 3,813.09 3,934.58 Pore water 3,726.75 4,119.99 Water Column Minimum 1,872.69 1,620.51 Pore water 1,111.80 2,239.71 Water Column Maximum 6,312.21 6,544.83 Pore water 7,081.23 5,595.60

Boron in both ponds has both pore water and water column samples that are both above and below the average seawater concentration in all sampling months. There is both a source and a sink of B in the pond systems through the entire growout season.

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Figure 140: TDS versus B for (a) Pond 7 and (b) Pond 10 (Blue dashed line 4.6 mg/L average seawater concentration from Hem (1989))

In Pond 7 the Spearman correlation coefficients highlight a very strong positive correlation between B and As (0.912) and B and Li (9.49) and weaker correlations between B and Cr (0.578), Rb (0.839), Sr (0.760) and V (0.852). In Pond 10 there are strong positive correlations between B and As (0.770), Li (0.978), Rb (0.811), Sr (0.599) and V (0.669).

11.6.2.2 Potential source minerals

Boron is essential in plant growth and so the farm management practice of adding fertiliser to the ponds to stimulate algal blooms is likely to provide excess B.

The depletion of Boron may relate to the biological production within the pond with algae, the aquaculture species and bacteria scavenging trace amounts of Boron to assist in building cell walls. Boron may also be complexing with various ligands to form insoluble complexes.

11.7 Radioactive elements

11.7.1 Uranium

Uranium is present in seawater in concentrations of about 3 μg/L. It is most common as 238uranium in oxygen rich environments, and bonds with oxygen to 2+ form the uranyl oxide UO2 . The uranyl ion complexes with carbonate and sulfate (Hem, 1989).

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11.7.1.1 Concentrations and Distribution in the ponds

The average concentrations for the water column and pore waters are presented in Table 35. The average concentrations in both the pore water and the water column are higher than the seawater concentration quoted by Hem (1989).

Table 35: Descriptive Statistics for Uranium in Pond 7and Pond 10 (Pond 7 n=105 and Pond 10 n=134)

Sample Location Descriptive Statistic Pond 7 Pond 10 Total Uranium Total Uranium (μg/L) (μg/L) Water Column Average 2.74 3.12 Pore water 1.58 4.21 Water Column Minimum 0.93 1.05 Pore water 0.21 0.60 Water Column Maximum 6.75 6.93 Pore water 4.80 29.19

Figure 141 shows the measured U concentrations: when compared to the average seawater U concentration the November samples are relatively enriched and the February, April and June samples are slightly depleted. This suggests that there was initially a U-bearing mineral in the ponds that was able to dissolve and increase the water’s U. After this initial dissolution phase (in the second and subsequent sampling periods) U was lost from the system, most likely as a precipitating mineral returning to the sediment as a solid.

Figure 141: pH versus UTOT for (a) Pond 7 and (b) Pond 10 (blue dashed line shows the average seawater concentration of 3 μμμg/L (Hem, 1989))

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In Pond 7 there are mild positive correlations between U and Ag (0.511), As (0.525), Cs (0.591), Mo (0.546) and Rb (0.583) with a weaker correlation between U and B (0.428), and U and Sr (0.429) in Pond 7. There are positive correlations in Pond 10 between U and As (0.624), Mo (0.511) Rb (0.508) and Zn (0.562) with a weaker correlation with U and Ag (0.416), and U and V (0.448) in Pond 10.

11.7.1.2 Potential source minerals

The spike in U at the commencement of the season is like to reflect the dissolution of a U-bearing mineral from the pond sediments. This mineral may have either: (1) accumulated in the pond sediments during the previous three months of atmospheric exposure (when the pond was emptied and prepared for the new season); or (2) was inadvertently added to the pond during pond preparation (potentially as an oxidised coating on the sand trucked in to line the ponds).

In either case once the initial dissolution was complete (probably through a reduction of a U-oxide) the pond conditions were able to remove the initial excess U, and then continued to provide a U-sink to the pond water through to the end of the grow-out season. This sink is most likely (given the low Eh in the pond water column and pore waters) to be in the water column or at the sediment –water interface through the precipitation of a U-carbonate.

11.8 Summary

The common theme from all the data presented in this chapter is that the trace elements tend to be more concentrated in the sediment pore waters than in the water column. For each element there are typically source and sink minerals that’s stability reacts with changing pond chemistry.

Figure 142 is a compilation of the data from both Ponds 7 and 10: it clearly shows that the greatest concentrations of trace elements (particularly in the case of iron) are associated with the pond pore waters. The pore waters are more acidic and more strongly reduced than any other part of the ponds and that makes them ideal for trace element liberation from minerals and mobility. Once in solution these trace elements move into the water column through the

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Figure 142: Average trace metal concentrations in the ponds at Tomei. Notice that the dominant trace elements in the sediment pore water were Fe2+ and Fe3+.

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12 HYDROGEOCHEMICAL PROCESSES

12.1 Introduction

The following is a list of dominant active processes in the water column and sediment pore water in the two prawn ponds studied at Tomei Pty Ltd:

• Mineral dissolution and precipitation

• Redox reactions

• Precipitation/dissolution reactions (influenced by redox reactions)

• Ion-exchange

• Mobilisation of metals

• Adsorption/desorption reactions

• pH buffering

• interaction between HCO3/CO3 (carbonate species)

• Photosynthesis and respiration

• Evaporatitive concentration

It is important to remember that these reactions are dynamic and some are reversible moving forward and backward on a daily basis.

12.2 Processes active in the pond base and dyke walls

12.2.1 Mineral dissolution and precipitation

12.2.1.1 Sodium mineral dissolution/precipitation

The dissolution/precipitation of halite (NaCl) most likely controlled the deposition of NaCl on and in the dyke walls and in the pond bottom sediment.

Pond wall material is likely to have been eroded during the use of paddle wheels in the ponds, from bioturbation in the pond bottom sediment or from the dissolution of salts (deposited on the pond walls) during a rainfall event. The dissolution of halite would have enabled the release of Na+ and Cl- ions into the pond water by the following reaction:

Equation 29 NaCl (s)  Na+ (aq) + Cl- (aq)

Although chemically unlikely, there is the possibility that during periods of high evaporation, salts precipitated out of the pond water and dyke pore water onto

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Chapter 12: Hydrogeochemical Processes the pond walls and banks. During rainfall events, the salts would have dissolved releasing bound metals, and would have been washed back into the pond water. (Hammarstrom et al., 2005).

12.2.1.2 Calcium mineral dissolution/precipitation

Calcium carbonate (CaCO3) dissolution was most likely contributing the greatest amount of Ca to the pond water. Seawater naturally contains Ca, however in the ponds studied; there was a lower concentration of dissolved Ca than in the reference seawater.

There is a loss of Ca from the pond water and this may be a function of reverse ion-exchange (Ca2+2Na+) or it may being scavenged by polychaetes in the pond and used for the construction of the CaCO3 worm burrows. It is questionable if the latter would effect the Ca concentration markedly, but looking at some of the ponds on the farm, there is an infestation of worms in some of the ponds as evidenced by the amount of worm burrows on top of the benthic zone (Plate 42). It is well known the macrobenthos with slightly acidic – slightly alkaline guts (pH range 5-8) are able to break down minerals such as

CaCO3 during digestive processes (Jansen and Ahrens, 2004). During digestion and movement through the gut, the minerals structure is mechanically and chemically impacted on. Metabolites and ingestive products such as mucus are able to chemically alter the pore water pH and redox conditions. Jansen and Ahrens (2004) agree that deposit feeders are able to play a significant (although possibly relatively small) part in biochemical cycles.

Understanding the bioavailability of Ca (and other elements) in seawater and how much is taken up by different aquatic organisms would be an interesting topic for further research.

12.2.1.3 Potassium mineral dissolution/precipitation

In Section 10.7 of this thesis it was established that both Ponds 7 and 10 had elevated average K concentrations compared to the seawater reference data. The dissolution of minerals such as feldspar, micas and K-rich clays through weathering reactions in the dyke walls and pond bottom clay sediment are most likely the dominant source of K ions in the pond water.

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12.2.1.4 Magnesium mineral dissolution/precipitation

Interpretation of Mg data from the two ponds studied in this thesis was discussed in Section 10.8. Findings concluded that Mg was not dissimilar to the reference seawater and therefore minimal to no precipitation/dissolution reactions were taking place. Slightly elevated concentrations of Mg in the pond water samples may be a result of small contributions of Mg in the form of dolomite (CaMgCO3) being added to the pond during pond preparation (liming).

12.2.1.5 Iron mineral dissolution/precipitation

Pyrite is the most likely source of iron mineral dissolution. XRD and previous studies (Gosavi, 2004) have identified there is a source of pyrite in the sediment at the farm. Pyrite is deposited in the ASS and undergoes dissolution over time. The oxidation of pyrite continues throughout the grow-out cycle and continues even after the ponds are drained. Pyrite contributes Fe and SO4 to the pond water. The Fe either stays dissolved in the pond water as its oxidised 3+ form Fe or precipitates as Fe(OH)3 (the red staining as seen on the dyke walls (Plate 2). It appears that a large amount of SO4 remains near the sediment-water interface where oxygen is low. As the SO4 is associated with a 2- reduced environment, it takes on the forms of S and H2S (g) or forms the black fine claylike material known as monosulfide.

12.2.1.6 Aluminium mineral dissolution/precipitation

Gibbsite (Al(OH)3) is most likely the dominant mineral to undergo dissolution in the pond water associated with ASS. This is because the dissolution kinetics of gibbsite is much more rapid than the other forms Al (such as the weathering of aluminosilicates) under acidic conditions (Appelo and Postma, 1999). The aluminosilicate weathering is more important for Al release over a period of years (rather than months) and this would most likely occur in the sediment pore water.

12.2.1.7 Manganese mineral dissolution/precipitation

The concentration of Mn is higher in the samples associated with the sediment pore water samples. Manganese can be introduced to the pond water by the dissolution of a Mn mineral (associated with igneous rocks), the

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Chapter 12: Hydrogeochemical Processes weathering/dissolution of Mn encrusted (adsorbed) to other mineral surfaces or the dissolution of a mineral that has Mn inclusions such as the ilmenite grains associated with the carboniferous sandstone deposited as bedrock at Pimpama (Preda and Cox, 2001). Therefore it is likely that Mn is introduced into the water column from the pore water sediment during the dissolution of Mn oxides, hydroxides and carbonates.

12.2.2 Redox (reversible reduction-oxidation) reactions

The common major elements that undergo redox processes are H, O, C, S, N, Fe and Mn. Some minor elements that have more than one oxidation state are U, Cr, As, Mo, V, Se, Sb, W, Cu, Au, Ag and Hg (Langmuir, 1997). From an environmental point of view, the oxidation state of an element is important as it dictates the elements toxicity and mobility. It is also important to remember that oxidation usually releases a proton (increases acidity) and reduction usually consumes protons (increasing the alkalinity) (Langmuir, 1997). Common redox reactions and their redox potential are displayed in Table 36.

Table 36: Redox reactions in pond bottom sediments (adapted from Reddy et al., 1986 and in Avnimelech and Ritvo, 2003).

Electron Acceptor Process Approximate redox (oxidising system) potential (mV)

O2  CO2 Aerobic respiration 500-600 - NO3  N2 Denitrification 300-400 Organic Components Fermentation <400 Fe3+  Fe2+ Reduction 200 Mn4+  Mn2+ 2- 2- SO4  S Sulfate reduction -100

CO2  CH4 Methanogensis -200

Ribet et al. (1995) described the process of facultative bacteria, living in organic-carbon-rich waters (like those in aquaculture ponds) using electron donors such as: O2, NO3, Mn(IV), Fe(lll) and SO4. These species assist in the oxidation of organic carbon for the bacteria’s energy source. The bacteria preferentially catalyse the most energetically charged species and work through the list (displayed in Table 36) when the most favourable element available is depleted.

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Piles of faecal material and organic matter on the pond sediment surface can some times cause anaerobic microcosms where sulfate reduction and denitrification may occur, surrounded by an oxic environment. A large proportion of the sediment oxygen uptake is not caused by aerobic respiration, but is rather due to re-oxidation of reduced inorganic metabolites (e.g. H2S) close to the oxic/anoxic interface. Up to 85% of the sulfide produced by sulfate reduction in marine sediments is not trapped permanently reacting with iron and other metals, but instead is continuously diffusing upwards to be re- oxidised in the surface sediment. About half of oxygen used by the sediment is consumed during the oxidation of sulfide in an indirect process with ions such as Fe, Mn and nitrate, or as a direct process (Kristensen, 2000).

Metal ions such as Fe (Fe2+ and Fe3+) and Mn have consistently higher concentrations in sediment pore water and are related to the sediment-water interface. This is due to low concentrations of dissolved oxygen and coupled with low Eh.

Both Ponds 7 and 10 have aerobic water columns and anaerobic sediment pore waters. There is a redox transition zone at the sediment-water interface.

The reactions listed in 12.2.2.1 are in order from the most favourable to the least favourable in relation to the amount of energy donated from the reaction to the bacteria (under particular conditions and pH 7 is assumed).

12.2.2.1 Reduction Reactions

The following reduction reactions take place in the sediment and affect the pond water column by the diffusion of pore water through the sediment-water interface into the water column. The reduction redox reactions are presented below in their stoicheiometric form (Freeze and Cherry, 1979):

Oxygen Reduction and the oxidation of organic material

This reaction occurs in the pond primarily at the sediment-water interface (Equation 30). This is where the decomposition of organic material such as uneaten feed, prawn excrement, and decaying aquatic fauna and flora occurs.

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Equation 30 CH2O + O2 (g)  CO2 (g) + H2O

A product of the aerobic respiration is CO2(g). The addition of CO2 into the pond water decreases the alkalinity of the water. This redox reaction supplies the most energy to bacteria and therefore is preferential as the first stage of oxidation in the pond.

Denitrification

When the concentration of dissolved oxygen is too small for use by microbes in the pond sediment, the oxygen atom bound in nitrate is used by microbes as their energy source for organic mater decomposition (Equation 31).

- - + Equation 31 CH2O + 4/5 NO3  2/5 N2 (g) + HCO3 + 1/5 H + 2/5 H2O

This process involves heterotrophic bacteria metabolising available carbon compounds and using nitrate as the electron acceptor when O2 isn’t present. - The nitrate is converted into N2(g) also with the production of HCO3 and + - proton acid (H ). The addition of HCO3 to the pond water increases the alkalinity and adds some inorganic carbon back to the pond water that was lost during nitrification. Denitrification reactions take place in the sediment, where there is a lack of oxygen due to the decomposition of organic material. The bacteria consume the organic matter using the nitrate as the oxygen source. This reaction would not happen in an oxidised environment, as the microbes would preferentially consume free oxygen because it takes less energy to - metabolise than the oxygen contained in the NO3 ion.

Fe3+ Reduction

Iron oxyhydroxides can be found in the sediment as mineral inclusions, as coatings on silicate grains, or as discrete grains. In the presence of organic matter and acidity the following reaction can occur (Equation 32).

+ 2+ - Equation 32 CH2O + Fe(OH)3 (s) + 7H  4Fe + HCO3 + 10H2O

The above reaction consumes a lot of proton acidity, however it produces reduced Fe (Fe2+) which, when released into an aerobic environment, bonds with oxygen quickly and precipitates as an iron oxide or hydroxide. The precipitation of iron oxyhydroxides occurs on the gill slits of respiring aquatic

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Chapter 12: Hydrogeochemical Processes organisms (where there is a source of O2 from the respiring animal) (Govinnage, 2001) and leads to the organisms eventual suffocation.

Manganese oxyhydroxides Reductive Dissolution

Most of the data collected from the three sampling periods at the farm exceed the ANZECC Guidelines (2000) for MnTOT. Data collected at the farm showed that the concentration of dissolved Mn was the greatest when pH was the lowest. Therefore the greatest concentrations of Mn occurred in the sediment pore water samples (where the pH was most acidic). Manganese minerals deposited in the sediment or as inclusions of ilmenite grains (found in Pimpama sediments by Preda and Cox, 2001) undergo dissolution in acidic environments, such as that found in the pond sediment pore water. Heavy metals have very high adsorption affinities for hydrated manganese oxides and during the dissolution of Mn minerals are released into the pond water. The following reactions (Equation 33 and Equation 34) can occur in this environment and are related to the reductive dissolution of Mn oxides and oxyhydroxides.

+ - + Equation 33 MnO2 + 4H + 2e  Mn2 +2H2O + - 2+ Equation 34 MnOOH + 3H + e  Mn + 2H2O

Mineral dissolution due to proton acidity, the presence of dissolved Mn in the pond water can be mediated by microbes or controlled by cation-exchange reactions. Figure 103 (in Section 11.2.2) shows that when DO decreases, MnTOT increases. This is most likely due to microbes using the oxygen from Mn oxides and oxyhydroxides to provide the energy for the consumption of organic matter and in doing this they, initiate the release of the Mn ion into the pore water. In the presence of bacteria and organic matter, the equation for manganese reduction would be as presented in Equation 35:

+ 2+ - Equation 35 CH2O + MnO2(s) + 3H = 2Mn + HCO3 + 2H2O

This reaction would be occurring in the benthic zone where there is a source of organic material (e.g. waste food), deposited manganese oxide and acid (from the other chemical reactions producing acid in the sediment zone).

Sulfate Reduction

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Sulfate reduction occurs when heterotrophic bacteria belonging to the Desulfovibrio or Desulfotomaculum genus consume organic matter to gain energy (Equation 36).

2- - - + Equation 36 CH2O + ½ SO4  ½ HS + HCO3 + ½ H

2- This reaction most likely occurs in the sediment pore water where SO4 is released from the oxidation of pyrite and where there is the greatest concentration of S2-.

During the reaction shown in Equation 36, the oxygen atom contained in the sulfate ion is used as the terminal electron acceptor and is consumed in the absence of free oxygen. The resulting products of the reaction are HS-

(generally in the form of H2S (g) as seen in, and smelt in samples taken from the sediment pore water at Tomei). This reaction facilitates the addition of proton acidity to the water and assists in decreasing the pH.

Iron that has been liberated from minerals such as pyrite, oxides, oxyhydroxides and clay minerals; reacts rapidly with the hydrogen sulfide to form amorphous iron sulfide and/or minerals such as greigite and mackinawite. When there are other trace metals present, there is a greater chance for the formation of trace metal sulfides. Iron sulfides have lower stability constants than the trace metal sulfides and therefore thermodynamically, the trace metal sulfides are likely to form before the iron sulfides. In saying this however, trace metals are also able to adsorb or co precipitate with iron sulfides during their formation.

Methanogenesis

Methanogenesis (or methane fermentation) is an anaerobic process whereby carbon is reduced to its most reduced oxidation state -4 in CH4 or methane (Equation 37).

Equation 37 CH2O + ½ CO2 (g)  ½ CH4 + CO2 (g)

When optimal redox conditions are reached, sulfate reduction and methanogenesis may occur almost simultaneously (Freeze and Cherry, 1979).

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Gas bubbles have been observed to be percolating from the organic sludge piles deposited on the bottom of ponds. It may be that they are a combination of H2S gas and CO2 gas.

12.2.2.2 Oxidation Reactions

Along with reduction reactions in the pond water, there are also reactions that involve the oxidation of elements. The oxidation reactions are presented below in their stoicheiometric form (Freeze and Cherry, 1979):

Sulfide oxidation

This process is likely to occur near the sediment-water interface where there is - + enough O2 to oxidise HS . In this case H is released into the pond water and assists in decreasing the pH.

- 2- + Equation 38 O2 + ½ HS  ½ SO4 + ½ H

This process can happen spontaneously (without the mediation of bacteria) in the presence of oxygen. When bacteria are involved, it generally is the Thiobacillus genus that assists in the chemical reaction. The reaction is thought to follow the steps:

Equation 39 ½ O2 + H2S  S° + H2O (where S° is elemental sulfur) 2- + Equation 40 O2 + S° + H2O  SO3 + 2H 2- 2- Equation 41 SO3 + ½ O2  SO4

Iron oxidation

Iron oxidation occurs at the sediment-water interface and at the surface of the pond where the pond dyke wall erosion is most prevalent (due to runoff and weathering). It also occurs where the maximum pyrite saturation indices are, that is, where pyrite is supersaturated and could precipitate under optimal conditions). In these areas of the pond, there is a source of Fe2+ (from the oxidation of pyrite) and acid, therefore the oxidised form of Fe is produced (Fe3+) (as in Equation 42) which in turn can bind with an oxide or hydroxide to precipitate as Fe(OH)3 on the dyke walls (Plate 2).

2+ + 3+ Equation 42 ¼ O2 + Fe + H  Fe + ½ H2O

Nitrification

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Nitrification can occur in the sediment or in the water column where there are - bacteria that oxidise ammonium to nitrate (NO3 ) and nitrite (NO2) (Equation 43).

+ - + Equation 43 O2 + ½ NH4  ½ NO3 + H + ½ H2O

Bacteria from the Nitrosomonous spp. and Nitrobacter spp. are involved in this reaction. The ammonium source in ponds is preliminarily from the waste of aquatic organisms in the form of ammonia excreted via the gills, by either simple diffusion or by sodium exchange mechanism (Dall et al., 1990) and the breakdown of aquaculture feed.

Manganese Oxidation

Manganese oxidation occurs where free Mn2+ comes into contact with oxygen and hydrogen atoms (Equation 44).

2+ + Equation 44 O2 + 2Mn + 2H2O  2MnO2(s) + 4H

It was proven in Section 11.2.2 that Mn was measured at the greatest concentrations when O2 was at its lowest. This suggests that in reduced conditions, such as in the sediment pore water, free Mn may precipitate once exposed to O2 (possibly due to bioturbation or microbiological activity).

Iron sulfide oxidation

Iron sulfide oxidation occurs when pyrite (contained in ASS) is oxidised and precipitates as a Fe oxyhydroxide (Fe(OH)3) (Equation 45).

2- + Equation 45 15/4 O2 + FeS2(s) + 7/2H2O  Fe(OH)3(s) + 2SO4 + 4H

The precipitated iron is most notably seen at the farm as the red stain on the pond bottom when ponds are drained and left to dry out; the red deposit on the walls of the settling drains (Plate 2); and the red stain beneath the geotextile on the pond walls (where reduced iron has leached out of the dyke walls and has oxidised once in contact with atmospheric oxygen).

12.2.3 Ion-exchange

Ion-exchange occurs when there is a shortfall or excess of ions through water- clay/colloid interaction. It is an electrostatic process, that is, lower charged

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Chapter 12: Hydrogeochemical Processes cations on the clays surface attract dissolved cations in solution to the surface of the clay to balance the net charge. Often ion-exchange is also often referred to as cation-exchange. The cation-exchange capacity (CEC) is dependent on the pH and the nature of the cations occupying exchange sites on mineral and surfaces. Organic matter and oxide precipitates which are likely to form coatings on clay’s and quartz grains, have very high cation-exchange capacities.

When cations with one charge (monovalent cations) are exchanged with cations with a double charge (divalent cation), the process is called reverse ion-exchange. The resulting equation is presented in Equation 46 (Drever, 1997):

Equation 46 2A-clay + C2+ = C-clay + 2A+

There was slightly more Na than Ca in the pond water samples (Section 10.5) and therefore ion-exchange is occurring between Na and Ca according to Equation 47:

Equation 47 Ca2+ 2Na+

The result was more Na in dissolution in the pond water than Ca.

12.2.4 Mobilisation of metals

Metals are mobilised when minerals (and clays): undergo dissolution reactions; are influenced by redox reactions; and, when they undergo desorption from mineral and clay surfaces. Concentrations of trace elements that exceed “normal” levels can be of environmental significance. When elevated they may contaminate surface and shallow groundwater and receiving estuaries. In addition, organisms and vegetation can uptake metals, increasing the potential for entry of some metals into the food chain.

When sampling in the natural environment it is important to remember that on first pass of the data some metals may appear to be enriched, however by comparing the naturally elevated metal concentrations to their source (e.g. bedrock material) it may be that the elevation is normal in that environment (Liaghati et al., 2003).

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Humic acid (organic material) can adsorb trace metals and release them at a later stage (therefore release of trace elements not related to mineral dissolution but to organics). Organic ligands can assist in the dissolution of minerals as they complex with metals (particularly Al) on surface sites and weaken the bonds between the metal and the solid. Trace metals can also be present as, adsorbed onto or inclusions in, other minerals and are released when the mineral dissolves (weathered).

Liaghati et al. (2003) suggested that adsorption and mobilisation of trace metals can be attributed to:

• influence of grain size

• influence of iron and manganese oxides

• occurrence and distribution of metals

Factors that control metal chemical behaviour are (Liaghati et al., 2003):

• sediment source (fluvial/estuarine)

• organic matter as organic carbon

• mineralogy (special regard to clay speciation)

• local land-use practices

• Iron hydroxide reduction mobilised trace elements such as As.

The dominant trace metals in the sediments of the Pumicestone region (southeast Qld) are Zn, V and Cr. Metals such as Mo, Co, Ni, Cu, Cd, Pb and As are common, but they are highly variable spatially (Preda and Cox, 2002). Preda and Cox (2002) stated that:

“all trace metals are controlled by the presence of Fe and Mn oxides, and by the grainsize of the sediment. Typically, fine-grained Fe-rich materials tend to absorb more trace metals than sandy sediments. In sediments that have developed from estuarine mud, some metals such as Cr, Mo, Pb and As tend to be in larger quantities than in the estuarine counterparts”.

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The dominant trace metals (metalloids) found in the ponds at Tomei are: As, V and Zn.

12.3 Process active at the sediment-water interface

The processes in the sediment-water interface are generally similar to those that are active in the pond walls; however there are a few key differences:

• mixing of the pond water leading to dilution;

• chemical buffering by high ionic strength seawater;

• changes in chemical potential of the environment related to pond water exchange;

• the high organic content associated with accumulation of feed, skeletal material, and faecal matter.

Generally, marine sediments follow a similar redox trend. The oxic zone is located at the top of a sediment profile. This is usually only a few centimetres thick (depending on bioturbation, water currents and benthic organism activity) and is dominated by the oxygen molecule. It is classified based on the depth of oxygen penetration. Below the oxic zone, is the suboxic zone. This region is characterised by ions such as nitrate, manganese oxides and iron oxyhydroxides. Beneath this, is the reduced zone. This zone contains ions such as sulfides which are produced during bacterial sulfur reduction. They are generally in either the precipitated form of iron sulfides or in the dissolved form as free sulfide. The black colouration in anoxic sediments is due to amorphous FeS (Allen et al., 1993).

Data collected from the ponds at Tomei and presented in this thesis highlight the strong geochemical differences between pore water and water column water. The sediment-water interface acts as the facilitator for many geochemical reactions occurring between the deposited ASS and the pond seawater.

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12.3.1 Mixing and dilution

12.3.1.1 Mixing

Mechanical agitation in the form of paddle wheels and aspirators were implemented at the farm to provide O2 to the pond water and to deter pond stratification. Before designing the sampling regime, the author tested pond water flow with the use of a flow meter. In Ponds 7 and 10 there was little to no movement of water near the sediment-water interface (tested by using a flow meter lowered down onto the base of the pond from the boat). During sampling periods the data confirmed that the mechanical aerators were not delivering acceptable amounts of O2 to the base of the pond, with the average DO concentration below 3.5 mg/L. Cultured prawns live in the benthic zone of the pond. Dissolved oxygen concentrations of <3.5 mg/L would have had a deleterious effect on prawn survival (reflected in the ponds poor productivity for the 2001/2002 season).

The paddlewheels were also found to be ineffective at mixing the water column. Sections of the pond contained less oxygen than other parts. Paddle wheels were turned off at night to save electricity, however pond stratification would have occurred and it may have been during these periods that Tomei suffered the majority of prawn mortalities. The data presented in Chapters 9, 10 and 11 suggest that the ponds were not being adequately mixed at times.

12.3.1.2 Dilution

All ponds on the farm received (on average) approximately a 10% water exchange each week. This meant that the ponds were refreshed with oxygenated seawater from the estuary. The only other mechanism for pond dilution was by rainwater. The salinity (EC) of the ponds studied highlighted the high evaporation rates in temperate climates, displayed in the samples collected in February. This data may suggest that the evaporation rate out- weighed the freshening effect of pond dilution. A solution to the problem of high salinity during the warm months may be to increase water exchanges. This would have assisted in maintaining optimal pond water salinities during periods of high evaporation and reduce the impact of increased salinity on the prawns.

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12.3.2 Chemical buffering by seawater

Seawater has a pH of 7.5-8.5 (Hem, 1989). It contains major cations (Na, Ca, - 2- - Mg and K) and major anions (Cl , SO4 and HCO3 ) from mineral weathering. The buffering capacity of seawater is its ability to accommodate the addition of an acid or a base without a notable change in pH (Stumm and Morgan, 1981).

Calcium carbonate in seawater has a neutralising capacity of about 2 moles/m3 (Dent and Bowman, 1996). Bicarbonate produced during denitrification, Fe3+ reduction, Mn oxide reduction and sulfate reduction reactions also assists in buffering acidic water released from ASS’s. It is also produced when CO2 is released from the pond sediment in the following reaction (Equation 48):

- + Equation 48 CO2 + H2O  HCO3 + H

The acid released from the underlying ASS in the ponds at Tomei was neutralised by the buffering capacity of seawater. Work carried out by Gossavi (2004) measured pH to be as low as 3.3 (on average) in the dyke pore water (with an EC of 31mS cm-1 - freshwater). However the lowest pH measured (on average) in Ponds 7 and 10 (in both the water column and pore water) was 7.7 (with an EC of about 52 000 mS cm-1 - seawater). This shows the capacity of seawater to neutralise the acid leached from ASS.

12.3.3 Adsorption/Desorption reactions

Adsorption reactions are very fast, much faster than dissolution and precipitation reactions. Adsorption is likely to occur when the pH of the pond water is alkaline. This process occurs on mineral grains (particularly on silicate (quartz grains) and clay surfaces). Ions such as Fe, Mn and Al are likely to be the dominant adsorbing ions; however inclusions such as trace metals can also be adsorbed onto mineral surfaces. This is usually due to surface and inner-sphere charges and cation-exchange reactions. The adsorption of some ions can occur even if they are only present in low concentrations in the pond water.

Adsorption can also occur with metals and organic matter through the complexation by chelators.

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12.3.4 Acidity /Alkalinity buffering

In any aquatic environment containing sediment, organic matter and microbes; biogeochemical reactions (involving the transfer of the H+ ion) result in a change in pH. These reactions occur relatively quickly and have the ability to change the pH of the pond water with such speed that the prawns do not have time to become accustomed to the change. Processes such as photosynthesis, denitrification and sulfate reduction increase pH, whereas respiration and nitrification decrease pH.

The pond water acidity at Tomei is influenced by: the oxidation of sulfidic material contained in ASS; by carbonate buffering through mineral dissolution; and, the dilution of pond water with water from the estuary by the farm managers and rainfall.

Changeable pH, particularly if the change is rapid towards acidic conditions, increases stress and potential mortality of the prawns. It also mitigates thermodynamic disequilibrium of the pond water chemistry. This may in turn manifest with rapid algal die back or “crashes” which will further decrease the pond pH. Minimising swings in pH between acidic and alkaline conditions will increase the stability of the water chemistry and in turn assure the good health of the aquaculture species.

12.3.5 Photosynthesis and respiration

Photosynthesis is the process whereby cells containing pigments such as chlorophyll a in plants and algae convert solar energy to organic matter (in the form of carbohydrates, fats, proteins, and nucleic acids) and oxygen. Respiration is the inverse of photosynthesis with carbon dioxide and water being the final products. Photosynthesis affects pH in the ponds as it uses CO2 during the days and respiration produces it during the night. During the day H+ is consumed and the pH rises, during the night CO2 is produced and the pH drops.

Equation 49 aaa Respiration

6CO2 + 6H2O aaababbb C6H12O6 + 6O2 bbb Photosynthesis

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Prawn farms promote the growth of algae as a supplementary food source and to provide shading for the prawns. A healthy pond is one that contains a dense (green) algal “bloom” in the water column and therefore the pond water is greatly impacted by photosynthesising algae. Photosynthesis and respiration are not restricted to the water column in the ponds studied. Aquatic flora growing in the benthic zone of the ponds would have affected the oxygen, carbon and nutrient concentrations at the sediment-water interface. Therefore, photosynthesis and respiration are important biochemical process affecting the ponds water chemistry.

12.4 Processes specific to the water column

12.4.1 Evaporation

The main source of salt in pond water is from the chemical weathering of minerals; however evaporation of the pond water also has an effect on the ponds EC. Increased EC has deleterious effects on the prawns and also the microalgae in the pond. Some species of microalgae will not grow when the EC is higher than that of natural seawater. Without the growth of microalgae in the pond water, the prawns suffer through a lack of UV shading and loss of nutritional feed.

12.4.2 Thermal stratification

As expected, the water temperature (measured from samples taken during the day) for both Ponds 7 and 10 are the highest in February (ranged between 26- 40 deg C). The other months had fairly similar ranges between (15-33 deg C). The paddle wheels assisted in destratifying the water column, however the paddle wheels are only turned on during day light hours and therefore thermal stratification would have occurred during the evening and possibly shown as a temperature lag until the paddle wheels mixed the entire pond.

12.5 Summary of the differences between the chemistry of the two ponds

There are subtle differences between the ponds; however, as they are both built into the ASS, they undergo similar hydrogeochemical processes. Pond 7 sediment is more acidic than Pond 10. The greater amount of acidity being

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Chapter 12: Hydrogeochemical Processes generated from the sediment of Pond 7 may be the reason why Pond 10 (production rate was 1,550kg) was more productive than Pond 7 (production rate was 53kg) for the 2001/2002 grow-out season (as they were stocked with similar amount of prawns at the beginning of the cycle about 33 prawns/m2).

12.6 Potential heavy metal affects on aquaculture species

The release of heavy metals to the aquatic environment is of concern. Once they are released from precipitated minerals, they are free to complex with organic matter, re-adsorb to other mineral surfaces, or remain in their toxic form to be bioaccumulated. The latter is particularly important as it is when absorption by aquatic organisms occurs and this can cause detrimental consequences to the organism. The author of this thesis intended to analyse tissue from dissected prawns from the farm to comment on the concentration of accumulated trace elements. However, there was a realisation by the author and the supervisors, that the analysis and interpretation of these data would have been extensive enough for a complete PhD thesis. Therefore a recommendation offered by the author (Chapter 13) is that future work could concentrate on this aspect alone.

Previous studies have concentrated on trace metals and the bioaccumulation by aquatic animals. There is some discussion, whether the toxic form (generally the reduced form) should be looked at rather than measuring a total concentration of the element. For this to occur, speciation methods have to be improved. One of the difficulties with this concept is that the location of the trace metal, depends on which form it is in. For example, when a trace element is in the gut of an animal (where the pH is acidic and the environment is reducing and there is little O2) the element may be extremely toxic, however one the trace metal is adsorbed, it might bond with oxygen and become less toxic.

Darmono and Denton (1990) undertook a study on heavy metal concentrations in Penaeus monodon (tiger prawn) and Penaeus merguiensis (banana prawn). They looked at 10 trace elements: Cu, Zn, Fe, Mn, Cd, Ag, Mg, Pb, Ni, Co and Hg in wild and farmed prawns. They collected prawns with net, sorted them into sizes, then killed and dissected the prawns to sort muscle tissue from gills

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Chapter 12: Hydrogeochemical Processes etc. and then froze them (-20°C). They carried out acid digestion on them with nitric acid on a hot plate (135 deg. C). They found that the muscle tissue contained lowest metal levels and the hepatopancreas contained highest levels of Zn, Fe, Cd. The gills contained a range of metals, however, Cu was greatest in gill tissue.

Vijayram and Geraldine (1996) analysed freshwater prawns for heavy metals. They dissected each prawn sample, separated the hepatopancreas, gill and muscle tissue, weighed and digested with concentrated HNO3 at 100°C in a Teflon destruction bomb. They used linear regression analysis and Students t test to test the significance of regression coefficients. All the prawn tissue metal concentrations are expressed as μg/g wet weight. The study indicated that the order of highest accumulation of heavy metals (Zn and Cd) in the freshwater prawns occurred in hepatopancreas > gill > muscle. They found that the hepatopancreas is the major detoxifying organ and its function appears to be like that of a sponge to “mop up” excess free metal from the blood.

Brown and Markich (2000) studied the effects of heavy metals on aquatic organisms. They found that dissolved metals in water disrupts the surface of the cell membrane. Secondly, the metal will sorb or complex itself to binding sites in the cell membrane surface and finally the metal is taken up through the cell membrane into the aquatic organism. They concluded that the metal must first come into contact with the surface of the cell membrane to gain a biological response (BR). The free metal ion is believed to be in rapid equilibrium with the cell surface binding sites. Those metals that are not are considered to be biologically inactive. Examples of these are colloidal metals and those metals complexed to strong organic ligands.

Without analysising the tissue from prawns collected from the ponds at Tomei, we are unable to determine if the prawns are uptaking heavy metals, however it is certain that there are “free” heavy metals in the pond environment for the prawns to ingest.

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Chapter 13: Conclusions

13 CONCLUSIONS

One of the most important things to remember when studying ASS’s and aquaculture pond is that they are chemically dynamic systems: that is they move backwards and forwards from chemical disequilibrium to equilibrium in relatively short periods of time. The minerals that are contained in the sediment alter, and are wholly reliant on the physical variables and climatic influences on the system.

The overall objectives of this thesis were outlined in Chapter 1. They were:

• To confirm whether the ponds studied were chemically homogeneous or heterogeneous

• If the ponds are heterogeneous, then to identify the spatial, temporal, physical and chemical characteristics in the water column (including pore water in sediments)

• To identify if the ASS contribute to the chemistry of the ponds

• To understand and build a framework to describe the relationship between chemistry in the pond water and soil

• To capture and collate data describing the pond environment

• To propose modifications to management practices to minimise the impact (if any) of ASS on the farms productivity

• Suggest ways of remediation and reclamation of ASS in similar environments

This thesis has satisfied the objectives listed above. When designing the methods for sample collection it was hypothisied that the ponds would be chemically stratified. Therefore a collection system was designed to measure subtle differences in pond water chemistry. The ponds studied were indeed chemically heterogeneous. This was shown in the interpretation of data in Chapters 8, 9, 10, 11 and 12. These chapters confirmed that both the sediment and the pond water were chemically and physically dynamic over time and space. The study showed that the sediment layer contained greater concentrations of cations, anions, nutrients, trace elements and heavy metals than the pond water. Along with elements contained in the farms runoff, the

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Chapter 13: Conclusions pond sediment layer and the pond walls contributed ions to the pond water column. The dissolution of ions into the pond water was facilitated by complex reactions mainly occurring at the sediment-water interface. The high amounts of nutrients in the pond (mainly from uneaten feed and prawn excrement) encouraged the proliferation of biomass (predominantly microalgae such as cyanobacteria and green algae). The algae rapidly altered the pond water environment from an oxidised state into a reducing one and back again during growth phases. When the algae are alive, they dominate the water column where they photosynthesise and respire, utalising dissolved nutrients and other trace elements in the water. As the biomass proliferates its productivity increases until it reaches a point where either there is a lack of nutrients for its continued synthesis or the physical environment becomes unfavourable (such as when it rains and there is a shift in salinity to the detriment of the algae). The pond environment becomes nutrient deficient and the algae go through a die back phase. During this time the algae sinks to the bottom of the pond causing eutrophication and resulting in a reducing environment. The environment encourages the proliferation of anaerobic bacteria that break down the dead algae. In turn, the underlying sediment is exposed to reduced conditions and the pH falls. Under the acidic conditions, precipitated elements dissolve and ions are released into the water column.

The other dominant chemical reaction occurring in the pond was the oxidation of the underlying ASS. Complex chemical reactions occur at the sediment- water interface due to the underlying ASS. The oxidised ASS releases acid (in the form of H+ ions), sulfur and iron. The acid released from the ASS moves through the sediment-water interface and dissolves deposited minerals. This mobilises trace elements that are previously locked up in the mineral sediment structures. The trace elements are released into the pond water column and are bioavailable to the organisms in the pond (including prawns). In the reducing environment at the bottom of the pond, the sulfur and the iron that are released from the ASS are present in their reduced form. That is the sulfur in the form of sulfide bonds with the free hydrogen ions to form hydrogen sulphide gas (which was present in the water samples collected from the sediment layers). The reduced iron would remain in a free Fe3+ state until it bonds with an anion or free radical for example oxygen. This precipitate was

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Chapter 13: Conclusions generally seen in the pond on the walls and on the bottom as a red iron oxyhydoxide.

Interpretation of chemical and physical data contained in this thesis significantly adds to the knowledge about chemical reactions that take place in aquaculture pond over time and space. With this knowledge, farm managers are able to better manage their pond by understanding the nutrient dynamics and the problems associated with deposited ASS. The following chapter section contains recommendations from this thesis for farm managers.

13.1 Farm Management

This study highlighted some management changes that could be made at the farm to improve water chemistry and prawn productivity. The following are observations made during the study and may assist in improving the productivity of the farm and as the farm at Pimpama is typical of many other intensive farms around the globe; the results from this thesis can be more broadly applied.

• Ponds are chemically heterogeneous not homogeneous. Therefore it is important to keep the water moving at all times. Switching off the aerators at night to save money on electricity will lead to oxygen deficiency at the sediment-water interface where the prawns are located and could lead to prawn mortality.

• The pond chemistry changes over time, therefore monitoring of the water chemistry by the farm manager before large fluctuations of chemistry occur is important (i.e. if the farm manager sees that the pH is decreasing over time, then they can add a buffering agent such as

NaHCO3 or a carbonate source to increase the pH to seawater levels of about 8 before the situation becomes too serious).

• Aerators did not deliver an adequate amount of oxygen to the sediment-water interface for use by the prawns. A greater amount of directional aspirators should be used in the pond to direct the oxygen towards the sediment (where it is needed and used by the prawns).

• The farm is potentially over feeding and there is no way to determine when the mortalities occur in each pond. Feed trays should be used at

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the farm to determine if overfeeding is occurring. If the ponds are being over-fed, the waste food will add to the nutrient load; bacteria will utilise the available oxygen at the sediment-water interface, leaving less available oxygen for use by the prawns. Anoxic conditions will form at the benthic zone of the pond and prawns will be forced to move to areas in the pond where there is more available oxygen. This causes stress and could lead to mortality. Efficient feeding will ensure a more manageable nutrient load at the bottom of the ponds.

• Nutrients can be added to the pond at the beginning of the grow-out season for promotion of algal growth but there needs to be a limit set for the pond to manage a chemically balanced system.

• The external environment affects pond chemistry. Seasonal affects will change the water chemistry of the ponds. Heavy rainfall will reduce salinity and could flush more ions such as trace elements and nutrients into the ponds. The water chemistry should be monitored directly after these events and the pond chemistry should be adjusted (with either chemical additives or water exchanges) to reduce the effects of these environmental influences. If possible, mitigation strategies should be put into place before seasonal fluctuations effect the water chemistry too dramatically.

• Focus sampling on the bottom of the pond where the aquaculture species live and not at the top of the water column. Many farm managers monitor the water chemistry by taking a sample from the surface of the pond. This study has highlighted the importance of monitoring the water chemistry where the prawns are living, as the ponds are heterogeneous (even with functioning paddle wheels).

• Prawn farmers (particularly those who are culturing burrowing prawns) may have better success at growing (water column favouring) fish in the ASS effected ponds rather than the sediment dwelling prawns. The water chemistry in the water column was much more resistant to rapid and unfavourable chemical fluctuations. In general, there are some points to remember when using ASS’s for this type of industry:

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13.1.1 Farm Practices that should be avoided in ASS

• Clearing of existing mangroves, as this increases the amount of PASS exposure.

• Undertaking freshwater aquaculture. Seawater buffers the ASS effects. pH would be exponentially more acidic if freshwater aquaculture was employed.

• Exposing un-oxidised acid sulfate soil to oxygen between grow-out cycles.

• Discharging acidic or metal-laden water into estuary.

• Stocking ponds with sediment dwelling production species. Species that live in the water column would not be as greatly impacted as the sediment dwelling species are.

13.1.2 Actions to avoid problems in aquaculture

• Identify and avoid acid sulfate soil affected land for new aquaculture development.

• Treat discharge water to remove mobile elements and acid to minimise effects to estuarine environment.

• Establish a chemical balance in ponds for greater production of the aquaculture species and a reduction in heavy metal leaching from ASS.

Aquaculture in acid sulfate soils increases risks to the environment and the economic sustainability of the farm.

13.2 Recommendations for further research The following points are recommendations for future work: • A study on microbial species present in the sediment would assist in determining which chemical reactions are mediated by microbes and which species of microbes promote these reactions. This knowledge may help in the bioremediation of ASS environments.

• A follow on project from the point listed above is to better understand the microbiology of sediment and water column species and how these relate to chemical reactions (attributed from ASS) – test sediments to see if they contain; or quantify the abundance of; microbes such as Thiobacillus spp. as they can increase the rate of inorganic oxidation

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reactions by 5-6 times. This might be beneficial in the breakdown of anoxic organic waste accumulated on the pond bottom.

• An isotope analysis of certain elements could be undertaken to determine if their elevated concentrations are from natural sources (associated with ASS) or from anthropogenic sources (from farming practices at Tomei and surrounding cane farms).

• Extension of Gosavi et al. (2004a) trace metal analysis of the pond macroalgae to determine if they are acting as a hold and release mechanism (hold metals when alive, release metals when they die).

• A project on the stratification of pond sediments at the bottom of a prawn pond on a smaller scale (1-2mm) than my project by using a multilevel piesometer (as peepers would not only be hard to install, disturb the sediment, but will also affect the (anaerobic) bacterial community on and in the sediment). This study would highlight the heterogeneity of the pond sediments and further the knowledge of biogeochemical reactions in this region.

• The author has undertaken a large literature review during this study, and it seems that the reasons behind algal blooms and crashes are not entirely understood. Therefore it is suggested that this should be a priority for further research given the importance of algal blooms to the early life stages of prawns and their role in controlling light penetration in ponds.

• Sulfur gas is documented to have been released from wetland soils, mangrove swamps and salt marshes (Devai and DeLaune, 1995). Could exposed areas of ASS be major sources of S release into the atmosphere and be causing pollution? More work needs to be done at prawn farms to determine if sulfur is being released from the pond water.

13.3 Final Comment

This study has highlighted the complexity of hydrochemical processes in aquaculture ponds. Many of these chemical processes directly affect the health of the aquaculture species. However, similar chemical processes occur

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Chapter 13: Conclusions in marine ASS environments and therefore findings from this study can be used as a foundation when studying similar systems.

By understanding the complex chemical reactions that are taking place in the pond we can better understand the effects that water chemistry has upon the aquaculture species. This is not only important in preservation of the natural environment and ensuring sustainable aquaculture, but is also imperative to the survival of the Australian aquaculture industry.

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Sayles, F.L., Wilson, T.R.S., Hume, D.N. and Mangelsdorf, P.C. (1973) In situ samplier for marine sedimentary pore waters: evidence for potassium depletion and calcium enrichment. Science 181: 154-156.

Schroeder, G.L. (1975) Night time material balance for oxygen in fish ponds receiving organic wastes. Bamidgeh 27: 65-74.

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15 APPENDICES

Appendix 1 – Weather Data

Table 1: Logan Weather Station (40854) rainfall data over the sampling period (www.bom.gov.au)

YEAR MONTH MONTHLY MONTHLY MONTHLY MONTHLY MEAN MAX MIN TOTAL (mm) (mm) (mm) (mm) 2001 Nov-01 4.9 40.6 0 148.2 2002 Feb-02 4.8 52.2 0 135.2 2002 Apr-02 2.9 27.2 0 86.6 2002 Jun-02 2.4 24.8 0 70.6

Table 2: Gold Coast Sea Way Weather Station (40764) rainfall data over the sampling period (www.bom.gov.au)

YEAR MONTH MONTHLY MONTHLY MONTHLY MONTHLY MEAN MAX MIN TOTAL (mm) (mm) (mm) (mm) 2001 Nov-01 3.9 34.0 0 116.2 2002 Feb-02 1.4 11.0 0 39.0 2002 Apr-02 2.9 42.0 0 86.4 2002 Jun-02 1.8 18.0 0 55.2

Table 3: Rocky Point Sugar Mill (40319) rainfall data over the sampling period (www.bom.gov.au)

YEAR MONTH MONTHLY MONTHLY MONTHLY MONTHLY MEAN MAX MIN TOTAL (mm) (mm) (mm) (mm) 2001 Nov-01 8.5 36.7 0 195.5 2002 Feb-02 2.7 16.4 0 32.4 2002 Apr-02 3.5 38.2 0 91.2 2002 Jun-02 3.3 21.8 0 76.0

Table 4: Logan Weather Station (40854) temperature data over the sampling period (www.bom.gov.au)

YEAR MONTH MONTHLY MONTHLY MONTHLY MEAN MAX MIN (Deg C) (Deg C) (Deg C) 2001 Nov-01 27.3 32.6 11.7 2002 Feb-02 30.3 35.4 16.4 2002 Apr-02 26.8 30.6 14.1 2002 Jun-02 22.1 28.2 4.0

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Table 5: Gold Coast Sea Way Weather Station (40764) rainfall data over the sampling period (www.bom.gov.au)

YEAR MONTH MONTHLY MONTHLY MONTHLY MEAN MAX MIN (Deg C) (Deg C) (Deg C) 2001 Nov-01 26.7 31.6 13.5 2002 Feb-02 30.4 35.0 18.6 2002 Apr-02 26.3 30.8 16.7 2002 Jun-02 22.1 26.7 7.0

Table 6: Logan Weather Station (40854) wind speed data over the sampling period (www.bom.gov.au)

YEAR MONTH WIND SPEED MONTHLY MONTHLY MONTHLY (km/h) MEAN MAX MIN (km/h) (km/h) (km/h) 2001 Nov-01 9 am 10.2 33.0 0.0 2001 Nov-01 3 pm 17.2 33.0 0.0 2002 Feb-02 9 am 12.5 24.0 2.0 2002 Feb-02 3 pm 21.0 44.0 5.0 2002 Apr-02 9 am 10.0 17.0 0.0 2002 Apr-02 3 pm 15.0 24.0 0.0 2002 Jun-02 9 am 8.0 24.0 0.0 2002 Jun-02 3 pm 14.3 33.0 0.0

Table 7: Gold Coast Sea Way Weather Station (40764) wind speed data over the sampling period (www.bom.gov.au)

YEAR MONTH WIND SPEED MONTHLY MONTHLY MONTHLY (km/h) MEAN (km/h) MAX MIN (km/h) (km/h) 2001 Nov-01 9 am 20.2 42.0 8.0 2001 Nov-01 3 pm 25.5 48.0 8.0 2002 Feb-02 9 am 19.8 39.0 8.0 2002 Feb-02 3 pm 27.3 52.0 11.0 2002 Apr-02 9 am 20.0 44.0 0.0 2002 Apr-02 3 pm 27.5 52.0 13.0 2002 Jun-02 9 am 19.0 52.0 5.0 2002 Jun-02 3 pm 22.1 46.0 5.0

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Appendix 2: Regional Lithological Summary Generalised Beenleigh Block (Basement) Lithologies

Cranfield et al. (1976) and Lohe (1980) provide descriptions of the lithologies that make up the Neranleigh-Fernvale Bed. These are summarised below:

Arenite

The grain size of this rock ranges from fine to medium-coarse, but are generally medium to coarse. The grains are subangular to subrounded and have been altered and metamorphosed by compaction. The feldspathic arenites contain plagioclase, altered orthoclase and microcline. The lithic fragments are composed of chert, schist and quartzite, with approximately 20% of the fragments being composed of sericite, chlorite, quartz and epidote, with zircon as an accessory mineral.

Greywacke

The greywacke in the Beenleigh Block are predominantly quartzo-feldspathic sandstone. When fresh, they are non-porous and green-grey in colour. When weathered they are friable and intersected by goethite veins. Greywacke is mined at Readymix Beenleigh Quarry (Peachey Rd, Luscombe, Qld). This site was purchased from Bell Resources in 1988 and now produces concrete, asphalt aggregates, manufactured sand, road base and flood rock.

Conglomerate

The conglomerate consists of rounded granitic-granodioritic, acid volcanic and minor limestone pebbles and boulders, with some shale fragments. The matrix ranges from coarse to very coarse sand.

Chert and Jasper

Chert is generally interbedded with shale and arenite. Chert ranges in colour from light brown to white, pale green and grey. In some cases, it is dark brown to dark red. This colour variation is interpreted by these authors to suggest that there is a secondary coating of iron oxide, along fracture and bedding surfaces. Massive chert units are located in the near-surface and surface areas and are Tertiary in age. Most frequently the jasper is found as poorly bedded or massive and is red in colour (Lohe, 1980).

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Shale

The shale is fine grained and interbedded with greywacke, siltstone and sandstone.

Intermediate to Basic Volcanic Units

The basic volcanic units are generally in the form of greenstone and metabasalt. These are light-dark green in colour and are fine grained (Lohe, 1980). They are interbedded with the chert and are the result of submarine eruptions e.g. Mt Tamborine pillow lavas.

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Appendix 3: Sediment Analysis Data Table 1: Sediment particle size data after sieving for 3 minutes on shaker with 60-63 micrometers sieve (BS410/1986 – stainless steel mesh and stainless steel frame) for Ponds 7 and 10.

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Table 2: LECO and LOI (loss on ignition) data from Pond 10 and 7 sediment core samples

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Table 3: Sediment Core – Laboratory data sheet containing sediment moisture content and sediment pH

Core # Total Interval Weight of Wet Dry Difference Depth pH Length (cm) Empty Weight Weight (Wet - Dry) of pH Profile (cm) Crucible (sample + (sample + Profile using crucible) crucible) (cm) glass using using door sml large large Mettler AE Mettler PL Mettler 260 1200 PL1200 Balance Balance Balance (24 hours) POND 7 2 33 0-10 84.23 290.30 243.73 46.57 4 7.72 8 7.85 10 6.50 10-20 86.80 309.14 254.73 54.41 12 6.30 14 6.18 16 6.29 18 6.40 20-30 87.20 312.71 262.27 50.44 22 6.01 24 6.09 26 6.14 28 6.09 30-33 87.75 138.09 128.12 9.97 32 6.28 2a 57.5 0-10 84.23 267.30 229.44 37.86 2 7.28 4 7.58 6 7.42 8 6.85 10-20 86.80 279.91 231.78 48.13 12 4.45 14 4.03 16 3.97 18 4.01 20-30 87.20 281.03 236.42 44.61 22 4.36 24 4.77 26 4.51 28 5.05 30-40 87.75 287.64 237.49 50.15 32 5.26 34 5.28 36 5.68 38 5.49

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Core # Total Interval Weight of Wet Dry Difference Depth pH Length (cm) Empty Weight Weight (Wet - Dry) of pH Profile (cm) Crucible (sample + (sample + Profile using crucible) crucible) (cm) glass using using door sml large large Mettler AE Mettler PL Mettler 260 1200 PL1200 Balance Balance Balance (24 hours) 40-50 85.81 302.50 248.77 53.73 42 5.81 44 5.84 46 5.84 48 5.86 50-57.5 87.67 302.15 247.75 54.40 52 6.20 54 6.00 56 6.05 4 48.5 0-10 84.23 249.54 193.53 56.01 2 7.95 4 7.87 6 7.62 8 7.34 10-20 86.80 276.79 217.66 59.13 12 7.18 14 7.25 16 7.12 18 7.31 20-30 87.20 299.39 233.06 66.33 22 7.15 24 7.12 26 6.97 28 7.03 30-40 87.75 349.12 279.75 69.37 32 7.12 34 7.13 36 7.18 38 7.28 40-48.5 85.81 257.67 217.51 40.16 42 7.11 44 7.03 48 7.80 5 26 0-10 88.38 246.51 207.40 39.11 2 7.70 4 7.31 6 7.10 8 6.89 10-20 89.16 345.15 285.33 59.82 12 6.94 14 6.93 16 7.00 18 6.92 20-26 84.51 228.01 194.54 33.47 22 6.85 24 6.96 26 7.28 7 40 0-10 84.23 203.41 2 7.63 4 7.16 6 7.32 8 7.04 10-20 86.79 264.35 12 7.11 14 7.11 16 6.88 18 7.19 20-30 87.20 300.39 22 6.87 24 6.97 26 6.90

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Core # Total Interval Weight of Wet Dry Difference Depth pH Length (cm) Empty Weight Weight (Wet - Dry) of pH Profile (cm) Crucible (sample + (sample + Profile using crucible) crucible) (cm) glass using using door sml large large Mettler AE Mettler PL Mettler 260 1200 PL1200 Balance Balance Balance (24 hours) 28 6.85 30-40 87.74 322.40 32 6.94 34 6.91 36 7.11 38 7.21 8 44.5 0-10 85.82 284.84 245.93 38.91 2 8.09 4 7.98 6 7.82 8 7.73 10-20 87.67 349.00 295.61 53.39 12 7.15 14 7.22 16 7.16 18 7.17 20-30 88.39 365.16 304.79 60.37 22 6.99 24 6.97 26 6.90 28 6.89 30-40 89.17 325.32 277.01 48.31 32 6.99 34 6.84 36 6.87 38 7.07 40-44.5 84.51 201.64 177.31 24.33 42 6.88 44 7.21 9 54.4 0-10 84.23 223.52 187.81 35.71 2 6.07 4 5.34 6 4.62 8 5.18 10-20 86.79 262.77 218.24 44.53 12 6.05 14 5.86 16 6.03 18 6.25 20-30 87.20 256.68 215.10 41.58 22 5.86 24 5.69 26 6.02 28 5.91 30-40 87.74 333.71 283.58 50.13 32 6.11 34 5.99 36 6.02 38 5.91 40-50 85.81 352.59 300.66 51.93 42 5.99 44 6.00 46 6.23 48 6.15 50-54.4 87.67 241.94 209.48 32.46 52 6.18 54 5.99 56 6.40 10 43 0-10 85.81 250.12 2 7.26 4 6.66

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Core # Total Interval Weight of Wet Dry Difference Depth pH Length (cm) Empty Weight Weight (Wet - Dry) of pH Profile (cm) Crucible (sample + (sample + Profile using crucible) crucible) (cm) glass using using door sml large large Mettler AE Mettler PL Mettler 260 1200 PL1200 Balance Balance Balance (24 hours) 6 6.23 8 6.32 10-20 87.67 295.94 12 6.05 14 5.84 16 5.48 18 5.31 20-30 88.38 353.79 22 5.54 24 5.63 26 5.52 28 5.53 30-43 89.16 435.21 32 5.56 34 5.59 36 5.76 38 6.03 42 6.11 11b 23 0-10 84.51 278.12 2 5.86 4 5.16 6 5.68 8 5.68 10-23 84.27 400.77 12 5.60 14 5.61 16 5.81 18 5.89 22 6.19 Core # Length Interval Weight of Wet Dry Difference Depth pH (cm) (cm) Empty Weight Weight (Wet - Dry) of pH Profile Crucible (sample + (sample + Profile using crucible) crucible) (cm) glass using using door sml large large Mettler AE Mettler PL Mettler 260 1200 PL1200 Balance Balance Balance (24 hours) POND 10 1 20 0-10 84.23 337.66 307.52 30.14 2 7.30 4 7.19 6 6.75 8 6.58 10-20 86.80 327.85 296.35 31.50 12 6.47 14 6.34 16 6.25 18 5.98 2 31 0-10 87.20 293.53 266.08 27.45 2 6.96 4 6.75 6 6.86 8 6.81 10-20 87.75 258.36 225.64 32.72 12 6.60 14 6.48 16 6.46 18 6.03

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Core # Total Interval Weight of Wet Dry Difference Depth pH Length (cm) Empty Weight Weight (Wet - Dry) of pH Profile (cm) Crucible (sample + (sample + Profile using crucible) crucible) (cm) glass using using door sml large large Mettler AE Mettler PL Mettler 260 1200 PL1200 Balance Balance Balance (24 hours) 20-30 85.81 259.17 220.21 38.96 22 5.95 24 6.11 26 5.89 28 4.10 3 16 0-10 87.67 269.99 242.03 27.96 2 6.58 4 6.24 6 6.05 8 6.19 10-20 88.38 193.92 174.76 19.16 12 6.38 14 6.25 16 6.44 4 22 0-10 84.23 297.06 257.15 39.91 2 7.21 4 7.09 6 6.88 8 6.97 10-20 86.79 313.61 261.33 52.28 12 6.05 14 6.22 16 4.64 18 4.05 20-30 87.20 125.27 117.27 8.00 22 5.72 5 32 0-10 87.74 296.10 256.05 40.05 2 4.47 4 3.47 6 4.38 8 5.65 10-20 85.81 264.58 223.68 40.90 12 5.29 14 4.63 16 4.35 18 4.25 20-30 87.67 243.20 201.52 41.68 22 4.25 24 4.76 26 4.82 28 5.15 30-32 88.38 117.39 110.59 6.80 32 4.51 6 14 0-10 89.16 322.33 285.50 36.83 2 6.02 4 6.58 6 6.86 8 5.46 10-20 84.51 169.24 147.84 21.40 12 3.99 7 31 0-10 84.23 267.07 2 6.83 4 7.08 6 6.64 8 6.11 10-20 86.79 261.66 12 5.66 14 5.83 16 5.87 18 5.97 20-30 87.20 263.46 22 5.69 24 4.91

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Core # Total Interval Weight of Wet Dry Difference Depth pH Length (cm) Empty Weight Weight (Wet - Dry) of pH Profile (cm) Crucible (sample + (sample + Profile using crucible) crucible) (cm) glass using using door sml large large Mettler AE Mettler PL Mettler 260 1200 PL1200 Balance Balance Balance (24 hours) 26 5.67 28 5.50 30-34 87.74 133.10 32 6.00 34 5.72 8 24.5 0-10 85.81 304.83 2 6.95 4 6.83 6 7.05 8 6.40 10-20 87.67 300.12 12 5.91 14 3.40 16 5.29 18 3.95 20-23.5 88.38 148.85 22 5.51 10 23 0-10 89.16 294.00 2 7.39 4 7.18 6 7.32 8 7.43 10-20 84.51 329.39 12 7.18 14 7.08 16 7.07 18 6.79 20-23 84.27 114.31 22 6.56 11 23 0-10 88.38 349.72 318.40 31.32 2 7.37 4 7.04 6 6.87 8 6.66 10-20 89.16 335.25 272.49 62.76 12 6.86 14 6.33 16 6.52 18 6.27 20-22.5 84.51 137.22 121.45 15.77 22 5.48

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Appendix 4: Water Chemistry Data

Due to the amount of water chemistry data collected during this study, the data can not be included in this Appendix and read clearly. Therefore data are contined on the CD at the back of this thesis.

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Appendix 5: Seawater Reference Data

Due to the amount of seawater reference data used during this study, the data can not be included in this Appendix and read clearly. Therefore data are contined on the CD at the back of this thesis.

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Appendix 6: Calculated Saturation Indicies

Due to the amount of calculated saturation indicie data used during this study, the data can not be included in this Appendix and read clearly. Therefore data are contined on the CD at the back of this thesis.

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Appendix 7: Spearman Correlation Coefficients

Due to the amount of Spearman correlation coefficient data produced during this study, the data can not be included in this Appendix and read clearly. Therefore data are contined on the CD at the back of this thesis.

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