Distribution and Speciation of Heavy Metals in Sediments from Lake Burragorang

by

Archana Saily Painuly

A Thesis Presented for the Degree of Masters of Engineering (Honours)

School of Engineering

College of Health & Science

University of Western Sydney

2006

1 Acknowledgments

At last I got this moment to write my gratitude to all those who have directly or indirectly supported me to make my long cherished dream to come to a reality.

At the outset, I express my profound sense of gratitude and respect to my chief supervisor Dr. Surendra Shrestha for his invaluable guidance and support academically as well as morally. Without his concrete suggestions it would have been impossible to bring out this work in the current form. Words are inadequate to express my heartfelt appreciation for Dr. Paul Hackney, my associate supervisor, who has been a constant help throughout this entire thesis. Special thanks to him to have helped me perform last stages of sampling during my third trimester.

I take this opportunity to thank Prof. Steven Riley, Head School of Engineering for providing me the infrastructural facilities to carry out this work. I am grateful to Technical staff of School of Engineering for constructing glove box and extrusion device for processing sediment samples.

I would also like to thank Professor Samuel Adeloju for arranging metal analysis in Australian Government Analytical Laboratories (Pymble, NSW). I am thankful to Dr. Honway Louie and Dr. Michael Wee and their staff at Australian Government Analytical Laboratories (Pymble, NSW) for performing metal analysis. . My sincere thanks to Dr. Henk Heijnis, Jennifer Harrison and Atun Zawadzki, from Environmental Radiochemistry at Australian Nuclear Science and Technology Organisation (Sydney, NSW) for their assistance with the 210Pb dating technique, determination of age profiles and interpretation of the data.

I would like to record my thanks to my work colleagues at Environmental Health Sharron, Sue, Burhan, Maree and Gavin for the moral support and encouragement. My thanks are due to Dr. Robert Mulley, Head of School Natural Sciences for approving my study leave.

I shall remain indebted to Dr. Arun Garg and Dr.Vinita garg for their moral support and valuable advice. Countless images flash through my mind when I remember the hard phase of time I was passing through, and here my husband, Nirmal, deserve a special mention who made the most conspicuous contribution in making this ambition reality.

I must also place on record my deep sense of love and tender sentiments for my family members for their perpetual encouragement and inspiration, despite being far away.

I will fail in my duty if I forget to mention ‘my bundle of joys’ Goura and Shriya, who were born during this period. They kept me cheerful even when the things were going tough.

2

Statement of Authentication

The work presented in this thesis is, to the best of my knowledge and belief, original except as acknowledged in the text. I hereby declare that I have not submitted this material, either in full or in part, for a degree at this or any other institution.

…………………………………………… (Signature)

3 Table of Contents

ABBREVIATIONS...... 9 ABSTRACT...... 10 CHAPTER I. INTRODUCTION ...... 12 1.1. BACKGROUND ...... 12 1.2. LAKE BURRAGORANG AND ITS CATCHMENT ...... 14 1.3. REPORT ORGANISATION...... 26 CHAPTER II. MATERIALS AND METHODS...... 28 2.1 FIELD SAMPLING...... 28 2.2 SEDIMENT GRAB ...... 28 2.3 SEDIMENT CORE...... 31 2.4 ANALYTICAL METHODS...... 33 2.4.1 MOISTURE CONTENT...... 33 2.4.2 ORGANIC MATTER AND CARBONATE CONTENT ...... 34 2.4.3 TOTAL NITROGEN AND PHOSPHORUS ...... 34 2.4.4 ACID EXTRACTABLE METAL...... 35 2.4.5 SPECIATION ...... 35 2.4.5.1 SEQUENTIAL EXTRACTION...... 35 2.4.5.2 SIMULTANEOUSLY EXTRACTED METAL (SEM) AND ACID VOLATILE SULPHIDE (AVS) ...... 36 2.4.6 SEDIMENTATION STUDY ...... 39 2.4.7 STATISTICAL TREATMENT OF DATA ...... 40 CHAPTER III. DISTRIBUTION OF METALS AND SPECIATION IN SEDIMENT OF LAKE BURRAGORANG USING SEQUENTIAL EXTRACTION ...... 42 3.1 INTRODUCTION...... 42 3.2 STUDY AREA...... 48 3.3 RESULTS AND DISCUSSION...... 48 3.3.1 METAL DISTRIBUTION ...... 48 3.3.2 METAL SPECIATION ...... 52 CHAPTER IV. DISTRIBUTION OF HEAVY METALS AND THEIR BIOAVAILABILITY USING SEM AND AVS IN THE SEDIMENTS OF LAKE BURRAGORANG ...... 61 4.1 INTRODUCTION...... 61 4.2 STUDY AREA...... 63 4.3 RESULTS AND DISCUSSION...... 64 4.3.1 ORGANIC MATTER AND CARBONATE CONTENT ...... 64 4.3.2 NUTRIENTS...... 65 4.3.3 BACKGROUND AND METAL DATA ...... 66 4.3.4 METALS...... 69 4.3.5 ACID VOLATILE SULPHIDE AND SIMULTANEOUSLY EXTRACTED METALS ...... 74 CHAPTER V. SEDIMENTARY RECORD OF HEAVY METAL POLLUTION OF LAKE BURRAGORANG USING 210PB DATING...... 78 5.1 INTRODUCTION...... 78 5.2 LEAD –210 RADIOMETRIC DATING...... 79 5.3 MODELS FOR SEDIMENTATION RATE DETERMINATION..81 5.4 SAMPLING LOCATIONS...... 81 5.5 SELECTION OF CORES...... 82 5.6 RESULTS AND DISCUSSION...... 82 5.6.1 CORE 1 (NEAR DAMWALL) ...... 82 5.6.2 CORE 2 (NEAR COX RIVER) ...... 83 5.6.3 CORE 3 (NEAR NATTAI RIVER)...... 83 CHAPTER VI. CONCLUSION ...... 90 REFERENCES ...... 95 APPENDIX A...... 114 APPENDIX B...... 115 APPENDIX C...... 119 APPENDIX D...... 124

5

List of Tables

Table 1.1. Warragamba catchment and its activities ...... 17 Table 2.1. Comparison of reference material values with obtained results...... 41 Table 3.1. Sediment quality guidelines for metals [Long et al., 1995]...... 49 Table 3.2. Metal distribution in the Lake Burragorang sediment grab samples according to sampling points...... 50 Table-3.3. Percentage of total metal content among the different sediment chemical fractions determined by sequential extractions ...... 53 Table 4.1. Lake Burragorang monitoring sites ...... 64 Table 4.2. Spatial and vertical distributions of carbonate content, organic matter and nutrients in sediment cores of Lake Burragorang ...... 67 Table 4.3. Variation in metal concentrations with depth in sediment core samples...... 71 Table 4.4. Background metal levels for Lake Burragorang from sedimentary metal concentrations ...... 73 Table 4.5. Background metal levels for Lake Burragorang with other matrices ...... 73 Table 4.6. Concentrations of AVS and SEM alongwith depth in sediments of Lake Burragorang ...... 75 Table 4.7. Guidelines for determining metal toxicity to benthic organisms in freshwater sediments (values in mg/kg) [Grabowski, 2001] ...... 76 Table 5.1. Activity variation of 210 Po, 226Ra and excess 210Pb with depth in sediment core 1...... 86 Table 5.2. Activity variation of 210 Po, 226Ra and excess 210Pb with depth in sediment core 2...... 86 Table 5.3. Activity variation of 210 Po, 226Ra and excess 210Pb with depth in sediment core 3...... 86 Table A-1 Uncertainty measurements for different studied variables ...... 114

6

List of Figures

Fig-1.1. Warragamba catchment showing Lake Burragorang [SCA, 1999] ..... 16 Fig-2.1. Locations of sediment core and grab samples in Lake Burragorang... 29 Fig 2.2. Ponar Petite sediment grab sampler...... 30 Fig 2.3. Sediment grab sample collected from Lake Burragorang...... 30 Fig 2.4. A) KB Sediment corer B) Sediment in an acrylic sediment core tube31 Fig 2.5. Sediment core extrusion device ...... 32 Fig 2.6. Top of sediment core stripper ...... 32 Fig 2.7. Details of sediment core stripper ...... 33 Fig 2.8. Flow chart of sequential extraction scheme for sediments metal speciation ...... 37 Fig 2.9. Extruding a sediment core in a glove box under nitrogen...... 38 Fig. 3.1. The concentration of metals in the sediment grabs from Lake Burragorang ...... 51 Fig 3.2. Metal distributions in Lake Burragorang sediments determined by sequential extractions ...... 55 Fig 3.3. Metal distributions in Lake Burragorang sediments determined by sequential extractions ...... 56 Fig 3.4. Metal distributions in Lake Burragorang sediments determined by sequential extractions ...... 57 Fig 3.5. Metal distributions in Lake Burragorang sediments determined by sequential extractions ...... 58 Fig 4.1. AVS and SEM distribution with depth ...... 77 Fig 5.1. Pathways by which 210Pb reaches lake sediments [Oldfield, 1981; Organo, 2000] ...... 80 Fig 5.2. Lake Burragorang Core 1 age versus 1) Rainfall 2) Metals 3) Organic matter and Carbonate content 4) Nutrients, Fe and Mn ...... 87 Fig 5.3. Lake Burragorang Core 2 age versus 1) Rainfall 2) Metals 3) Organic matter and Carbonate content 4) Nutrients, Fe and Mn ...... 88 Fig 5.4. Lake Burragorang Core 3 age versus 1) Rainfall 2) Metals 3) Organic matter and carbonate content 4) Nutrients, Fe and Mn...... 89

7 Fig B-1. Depth distributions of carbonate content, organic matter and nutrients in sediments...... 115 Fig B-2. Depth distributions of carbonate content, organic matter and nutrients in sediments...... 116 Fig B-3. Depth distributions of carbonate content, organic matter and nutrients in sediments...... 117 Fig B-4. Depth distributions of carbonate content, organic matter and nutrients in sediments...... 118 Fig C-1. Depth profiles of metals in sediments...... 119 Fig C-2. Depth profiles of metals in sediments...... 120 Fig C-3. Depth profiles of metals in sediments ...... 121 FigC-4. Depth profiles of metals in sediments...... 122 Fig C-5. Depth profiles of metals in sediments...... 123 Fig D-1. Core 1 profile of A) Po210 B) Ra210 C) excess Pb210 activity and D) age versus depth...... 124 Fig D-2. Core 2 profile of E) Po210 F) Ra210 G) excess Pb210 activity and H) age I) excess Pb210 activity normalised with <63 μm size versus depth ...... 125 Fig D-3. Core 1 profile of A) Po210 B) Ra210 C) excess Pb210 activity and D) age versus depth...... 126

8

Abbreviations

ANZECC Australian and New Zealand Environment and Conservation Council AVS Acid volatile sulphide As Arsenic AWT Australian water technology BL Background levels Cd Cadmium Cr Chromium CIC Constant initial concentration Co Cobalt CSIRO Commonwealth scientific and industrial research organisation Cu Copper ERL Effects range-low ERM Effects range-median Fe Iron HM Heavy Metals Hg Mercury ISQG Interim sediment quality guidelines Mn Manganese Mo Molybdenum Ni Nickel Pb Lead Po Polonium Ra Radium SCA Sydney catchment authority Se Selenium SEM Simultaneously extracted metals SOI Southern Oscillation Index STP Sewage treatment plant TN Total nitrogen TP Total phosphorus USEPA United states environmental protection authority V Vanadium Zn Zinc

9

Abstract

Lake Burragorang, the focus of this thesis, is the main water supply source for the large population of Sydney and is a major source for the Blue Mountains residents. This study was aimed to evaluate the distribution of heavy metals and their speciation in sediments of Lake Burragorang. The principal focus is on the study of heavy metal pollution and their bioavailability to the aquatic system.

Sediment grabs and core samples were collected and analysed for the determination of As, Cd, Cr, Co, Cu, Fe, Pb, Mn, Hg, Mo, Ni, Se, V and Zn. Based on the analysis, background concentrations were established as 4.7, 0.2, 23, 12, 20, 29000, 22, 660, < 0.2, 0.25, 19.7, 0.13, 37 and 68 mg/kg for As, Cd, Cr, Co, Cu, Fe, Pb, Mn, Hg, Mo, Ni, Se, V and Zn, respectively. Concentration of Hg and Se in all locations except at the sites DWA3 and DWA2 (refer Fig. 2.1 for location details) were found below the detection limits (0.1 mg/kg). The metal concentration was found to decreases in the order Fe > Mn > Zn > V > Cr > Pb ≅ Ni ≅ Cu > Co > As> Mo> Se > Cd. Overall metal distribution picture depicted that locations close to the dam wall had higher pollution compared to the other sites.

A five-step sequential extraction procedure was employed to assess different geochemical forms of these metals in sediment grabs of lake Burragorang. This is the first study to report metal speciation data for lake Burragorang sediments. No significant spatial variations were observed in the speciation trends. Hg and Se were not considered for speciation due to their low concentration observed in lake sediments. Substantial amount of metals like Cd, Co, Mn, and Zn were present in the first three fractions exchangeable, carbonate and reducible. The total Fe in the sediments is quite high which is alarming since its presence even in small amounts bound to the exchangeable and carbonate fraction could cause deleterious effects. The results showed the ease with which metals leach from sediments, decreases in the order: Mn=Cd>Co=Zn>Ni>Mo>Pb>Fe>V>As>Cu>Cr.

Sediment cores collected from various locations of Lake Burragorang were analysed for organic matter, carbonate contents, nutrients and metal concentration to

10 understand the history of pollution events that have occurred over an extended time span. Acid volatile sulphide and simultaneously extracted metal experiments were conducted on selected cores to have a better understanding of bioavailability of metals (usually Cd, Cu, Ni, Pb and Zn).

Total phosphorus (TP) ranged from 60mg/kg at UWS13 to 1360 mg/kg at DWA2 and total nitrogen (TN) ranged from 314mg/kg at DWA18 to 3769mg/kg at UWS14. The concentrations were generally higher at the top and decreased with depth.

The study showed that dominant metals were Fe and Mn followed by Zn, V, Cr, Pb, Ni, Cu, Co and As. Other metals such as Cd, Mo and Se were present in lesser amounts and, at few sites, were closer to the detection limit. Mercury was below detection limit in all locations. The highest sulphide levels were obtained from site

DWA2 (ranged from 0.59 to 0.12 μmol/g), while lowest levels were obtained from site DWA35 (ranged from 0.25 to 0.09 μmol/g). No regular trend was observed in the AVS (Acid Volatile Sulphide) pattern of the cores. In all the sites among HCl- extractable metals (SEM), the Cd concentrations were the lowest and the Zn was the highest. The results showed that these simultaneously extracted metals at all stations were higher than AVS and ratio was found greater than 1, which indicated that available AVS is not sufficient to bind with the extracted metals. This revealed that AVS is not a major metal binding component for Lake Burragorang sediment and contained metals, which could be potentially bioavailable to benthic organisms.

Sedimentation rates and age profiles on few preselected locations of Lake Burragorang were estimated using 210Pb dating method as described by Brugam [1978]. The variation in metals and nutrients in the sediments with age was established and has been compared with published historical record, rainfall records and bushfire data. Two cores from riverine zone (DWA18 and DWA35) and one from lacustrine zone (DWA2) were selected to perform sedimentation rate study using 210Pb dating method. The sedimentation rates for core 1, core 2 and core3 were calculated to be 0.47 ± 0.07, 0.19±0.004, 0.43±0.09 (cm/year), respectively. The ages calculated were used to establish the 50-year geochronology of changes in organic matter, carbonate content, nutrients and metal concentrations. Correlation was made up to 25 cm depth in core 1 and 3, and 15 cm depth in core 2 as cores demonstrated a decay profile up to these depths only. 11 Chapter I. Introduction

1.1. Background

Surface waters, including streams, rivers, natural and man-made lakes and oceans, are the support medium for most life on Earth. This life includes humans, who take most of their drinking water from surface systems. Unfortunately, humans have allowed surface waters to be the prime repository of our wastes – wastes from our bodies, our activities, and our great variety of conveniences and facilities (including manufacturing plants). In the pollution study of aquatic systems, heavy metal pollution assumes great significance. Metals constitute an important group of environmentally hazardous substances, some of which prove to be harmful to the very life that depends on the receiving water. The primary stress is toxicity to aquatic plants and animal organisms, but we are now very familiar with several secondary impacts; for example, bioaccumulation and bioconcentration of chemicals through the food chain that results in toxicity to non-aquatic species [Allen et al., 1997].

In the environmental community the notation of heavy metals implies stable high- density metals (lead, cadmium, mercury, copper, nickel etc.) and some metalloids (e.g. arsenic etc) [Ilyin, 2003]. The metals that referred to as heavy metals comprise a block of all the metals in Groups 3 to 16 that are in periods 4 and greater of periodic table [Hawkes, 1997].

Metal gain access to aquatic environment by natural process viz, weathering of soil and rocks, volcanic eruptions, and major transportation from terrestrial sources under high runoff from storms and floods. In addition, discharges from urban, industrial, mining and other human activities are other potential sources of particulates. The majority of heavy metals and their compounds possess pronounced properties of toxicants [Allen et al., 1995; Wright and Mason, 1999].

The accumulation of heavy metals in the bottom sediments of water bodies and the remobilization of these substances from the latter are two of the most important mechanisms in the regulation of pollutant concentrations in an aquatic environment [Linnik and Zubenko, 2000]. In the past, however, water quality studies focused 12 mostly on the detection of contaminants in the water column and ignored the fact that sediments may act as large sinks or reservoirs of contamination [Horowitz, 1991; Loring, 1991; USEPA, 2000]. Many past studies also failed to recognise that remobilization of metals from contaminated sediments can cause water quality problems [USEPA, 1999].

The heavy metals (HM) pollution of aquatic ecosystems is often most obviously reflected in high metal levels in sediments, macrophytes and benthic animals, than in elevated concentrations in water. The ecological effects of HM in aquatic ecosystems and their bioavailability and toxicity are closely related to species distributions in the solid and liquid phases of water bodies [Linnik, 2000]. Unlike the organic pollutants, heavy metals are not removed by natural processes of decomposition. On the contrary, they may be enriched by organisms (biomagnification) and can be converted to organic complexes which may be more toxic [Forstner and Muller, 1973]. They are always present in aquatic ecosystems and redistribute only among different components. This phenomenon has both positive and negative features.

While the bottom sediments promote self-purification in the aquatic environment because of HM accumulation, under certain conditions the bottom sediments can be a strong source of secondary water pollution [Denisova et al., 1989; Linnik et al., 1993]. The release of HM from bottom sediments is promoted, for example, by a deficit in dissolved oxygen, a decrease in pH and redox-potential (Eh), an increase in mineralisation and in dissolved organic matter (DOM) concentration. The mobility of HM depends on their forms of occurrence in the solid substrates and pore solutions of the bottom sediments, as well as on the physico-chemical conditions that arise on the boundary of solid and liquid phases, as noted previously. HM flow from pore solutions is one of the most important ways of exchange between bottom sediments and water [Linnik and Zubenko, 2000].

The specific toxicity mechanism of each metal is influenced by its characteristics, namely molecular configuration, solubility, particle size and other physico-chemical characteristics. The total concentration of a metal is determined for most environmental studies. This is a valid approach when studying mass balance. Total metal concentration is only helpful to identify change due to different possible 13 phenomena such as erosion, climate variability and leaching to groundwater. However, when the reason for a study relates to fate and effects, knowledge of the physico-chemical forms (i.e. species) is required. Metal speciation has become an important area of concern because of its importance in the understanding of the fate and effects of metals in the environment [Kramer and Allen, 1991].

The chemical properties and behaviour of these metal pollutants influence their fate, exposure and toxicity. The primary determinant of behaviour is the chemical form in which the metal occurs – referred to as the species of the metal. Metal speciation is therefore defined as the process or combination of processes by which a metal arrives at the form(s) in which it is found in a particular state of the environment, often the equilibrium state. Speciation can also rather loosely refer to the analytical determination of the species present in a particular state. Valence changes of the metal atom, the formation of oxyanions, complexation with inorganic or organic ligands in solution, sorption to particulate or sedimentary matter, precipitation, and interaction with microbes are among the processes that lead to a new distribution of metal species [Allen et al., 1997].

The present study was aimed at studying the distribution of heavy metals and their speciation in sediments of Lake Burragorang. High water quality from this lake is of crucial concern as it accomplishes the need of drinking water for over 4 million people of Sydney. Lake Burragorang’s inflow has a large range of water quality, which enters the lake from the six major tributaries. Water quality has been poor in Lake Burragorang during wet years compared to dry years as a result of pollutants and nutrient loading from the catchment. Sydney Catchment Authority reported elevated levels of phosphorus, nitrogen, iron, aluminium and manganese in lake water [SCA, 2001a]. There are number of activities within the catchment which could potentially pose a risk of metal pollution to water and sediment quality of the lake. The following section will describe the study area and its major activities in details.

1.2. Lake Burragorang and its Catchment

Lake Burragorang in south west of Sydney (Fig. 1.1), impounded by Warragamba Dam, is the main source of water supply for Sydney and is a major source for the 14 Blue Mountains. It provides approximately 80 per cent of the water to a population of about four million people. Lake Burragorang is one of the largest domestic water supply storages in the world, holding 2,057,000 million liters of water [SCA, 1999].

The lake is fed by several major rivers (Fig.2.1). The Wollondilly, Nattai, Kowmung and Coxs Rivers supply approximately 83 per cent of the total inflow to the lake. The waters within the lake and the Kowmung River are classified Class S – Specially Protected Waters. All other inflows are Class P – Protected Waters. These classifications reflect the significance of the storage and its tributaries for water supply purposes. The catchments of these rivers have differing geological, topographic and land use characteristics, which result in contributions of varying water quality to Lake Burragorang. These river systems rise outside the Warragamba Special Area (which consists of the stored waters of Lake Burragorang and adjacent lands), hence their water quality and that of Lake Burragorang is influenced by activities in the outer catchment areas [SCA, 1999]. In order to understand the water and sediment quality of the lake and their possible sources, it is prudent to discuss the surrounding areas and activities in these areas.

Warragamba catchment covers an area of approximately 905,000 hectare (ha) and is divided into two zones. The inner zone or catchment (or special areas) covers approximately 258,400 ha, comprises about 28 per cent of the total hydrological catchment and consists of the stored waters of Lake Burragorang and adjacent lands. It extends from the township of Warragamba in the northeast, to Buxton in the southeast, Wombeyan Caves in the southwest and to Narrowneck and the Wild Dog Mountains in the northwest. The remaining 72 per cent of the Warragamba hydrological catchment is known as ‘the outer catchment area’ (Figure 1.1) and includes the regional centers of Goulburn, Lithgow, Bowral, Mittagong, Katoomba and parts of the Blue Mountains townships of Mount Victoria, Blackheath, Leura and Wentworth Falls [SCA, 1999].

Warragamba catchment and its activities are summarised in Table1.1. The Coxs River catchment is located northwest of Sydney and covers an area of 2630 square kilometers, which includes the major urban areas of Lithgow and the southern edges of Katoomba. Coxs River catchment comprises 31% of the total catchment of the lake [Fredericks, 1994; Siaka, 1998]. The Coxs River catchment supplies up to 30% 15 of the water that is stored in Lake Burragorang. Siaka [1998] comprehensively described the Coxs catchment and its activities in his Masters thesis.

Fig-1.1. Warragamba catchment showing Lake Burragorang [SCA, 1999]

16 Table 1.1. Warragamba catchment and its activities

Warragamba Area Total Inflow Major Major activities subcatchment (Sq area (%) urban contributing to Km) (%) areas pollution

Power station, STP, Coal Lithgow, mines, Other small Coxs 2630 29 30 South industries -Copper ore Katoomba Refining, Pottery, Brick and Pipe works

STP, Ceased coalmines, Mittagong, Swimming pool, Nattai 369.1 4 11.5 Bowral, Discharge from industrial Mossvale and urban runoff

Agriculture, Grazing, Goulburn, STP, Pig and Poultry Wollondilly 3403 37.6 41 Marulan enterprises,Stables, Meat and wool processing

Unsewered residential Oaks, development, Werri Berri 160 2 0.5 Oakdale,W Agriculture, Livestocks, allacia Vegetable growing and Poultry farming

Several rivers join Coxs River before it enters Lake Burragorang. Main tributaries include Kowmung, Jenolan and Kedumba Rivers in the lower catchment. Pipers Flat, Marrangaroo and Farmers creek in the upper catchment and River Lett, Little River and Megalong creek in the middle catchment. Lake Wallace and Lake Lyell located at Wallerawang and south west of Lithgow city, respectively impound the water of the Coxs River before it flows to Lake Burragorang [Organo, 2000].

Mt. Piper and Wallerawang power stations are located in the upper catchment and they provide New South Wales (NSW) with approximately 30% of its coal- generated electricity. Waste ash resulting from coal burning is a potential source of water pollution, particularly when the ash is disposed of in landfill ash dams. Treated wastewater from Wallerawang power station including water used in

17 cooling towers and runoff from coal stockpiles at both power stations discharges directly into the upper Coxs River catchment [NSWEPA, 1993]. Sewage treatment plants (STPs) are other significant contributor to pollution in the catchment. STP typically releases pollutants including chlorides, oxygen-demanding (organic) wastes, ammonia and metals such as Pb and Cd [NSWEPA, 1993]. Several of these older plants in this area were not designed to remove nutrients from the water they discharged. Elevated levels of nutrients, nitrogen and phosphorus, in treated sewage effluent can cause problems in receiving waters because they lead to excessive growth of algae, floating weeds and attached plants. Therefore, water quality can be impaired by the production of objectionable odours and tastes, clogging of waterways can occur and consequently decreases the use of the waterway as a recreational amenity [Siaka, 1998].

The Kedumba River is a major tributary of the Coxs River and has high levels of nutrients and suspended solids. Elevated levels of algae are found in Lake Burragorang, where the Coxs and Kedumba Rivers enter in to the lake. South Katoomba sewage treatment plant and urban runoff from the Katoomba area are the likely sources of these levels in the Kedumba River. South Katoomba sewage treatment plant was closed in April 1998, following diversion of sewage flow to the Blue Mountains sewage transfer tunnel. Urban runoff from the Katoomba area will continue to contribute nutrients and suspended solids to Lake Burragorang during wet weather. Mount Victoria sewage treatment plant, operated by Sydney Water Corporation, and the Lithgow and Wallerawang sewage treatment plants, operated by the local Councils, release effluent into tributaries of the Coxs River [SCA, 1999]

Heavy metals and chemicals enter Coxs River directly and via its tributaries – Ncubeeks Creek carries Mn and Al, while Sawyers Swamp Creek carries Se, Mn, Fe, B, F, As and Sb. [CSIRO, 1990] reported increased concentrations of Mn, Zn and P in the sediment of Lake Wallace between 1985 and 1989.

Other possible sources of pollution are coal mines. The major impacts from coal mine operations are acid mine drainage, salinity and sedimentation. Most coals contain many trace elements, some of which have concentrations up to 1,000 mg/kg [Swaine, 1990]. The New South Wales EPA noted that water quality in the waterways of the catchment around Wallerawang, Lithgow and Hartley had 18 deteriorated as a result of the practices of open-cut coal mining operations being conducted in the bed, or on the banks of rivers and tributaries [NSWEPA, 1993].

Besides coals, the mining of base and precious metals is an important activity in the catchment [Maidment, 1991]. Swaine [1990] had analysed coal samples and found that the concentrations of Cd, Cr, Co, Cu, Pb, Mn, Ni and Zn in the coal were 0.062- 0.33, 3-30, < 2-10, 2-50, 2-24, 2-800, < 5-50 and 10-15 mg/kg, respectively. The study reported that in the surface mining of coal, some of these trace elements might have mobilised, especially under oxidising conditions. Consequently, this might have caused changes in the concentrations of some elements in nearby waters.

Lithgow City is the largest urban center within this catchment and provides many commercial services and supports a range of light industries. Heavy metals in Farmers Creek sediments are derived from natural sources and human activities, including copper ore refining and the operation of a blast furnace for producing pig iron and steel [Cremin, 1987]. Other human activities which may release heavy metals into the environment include coal mines, pottery works, brick works, pipe works, a small arms factory, extensive railway activities over a long period and a large number of cars and trucks driving through, or near Lithgow. The refining of Cu ores which contained typically 17.75% Fe, 14.5% Zn and 8.04% Pb [Crane, 1988] probably contaminated the environment [Siaka, 1998].

The Nattai River is a major sub-catchment of Lake Burragorang [McCotter, 1996; AWT, 2001]. The catchment covers an area of 369.1 Km2 in the southeast sector of the Burragorang catchment, which is approximately 3% of the total for the Burragorang water supply catchment [Anon, 2000]. Nattai originates near Mittagong 150km southwest of Sydney, and flows in a northerly direction for approximately 80km before entering the eastern shore of Lake Burragorang [Sydney Water, 1993]. The main sources of pollution to the Nattai catchment, highlighted in the reports by McClellan [1998] and [Anon [2000] include the Mittagong STP, the Welby Waste Disposal Area, and local settlement and industrial areas, namely Hill Top, Colo Vale and Mittagong. These causes have been identified for decline in water quality within the headwaters of the Nattai River [AWT, 2001].

19 The Nattai River is the steepest of all streams feeding the Warragamba dam storage. It is polluted by treated sewage, which discharges into it from Iron mines Creek, and by urban runoff from Mittagong. Iron mines Creek has turned cloudy brown in colour because of Mittagong's discharges. Excessive weed growth and a drain-like smell are apparent in upper parts of the Nattai River. Sediment associated with urban runoff provide a suitable substrate for weed establishment. The Nattai, being so short and steep, does not have much pollution absorption capacity, and such pollutants find their way into Lake Burragorang. Such continuing pollution severely degrades the Nattai River as well as Sydney's main water supply. Seepage and storm water runoff from the Welby Tip may also pollute the Nattai River with plant nutrients, heavy metals and other toxic substances. The tip is also a source of weed infestation and possibly of plant pathogens such as Phytophthora cinnamoni [Anon, 1999].

Chlorinated water discharge from the Mittagong Swimming Pool; storm water discharge from industrial and urban areas in both Mittagong and surrounding settlements (eg. Colo Vale, Hilltop), heavy metal, hydrocarbon and debris associated with the Freeway and Great Southern Railway are other major sources contributing to deteriorating effects on the quality of water in the Nattai River [Anon, 2000].

There has been a relatively long history of mining around the Nattai wilderness. In the west, mining of silver and lead ore at Yerranderie commenced in 1897. The town was home to over 2,000 miners by 1911. However, the boom was short-lived as the mine ceased to operate commercially by 1925, and was finally closed down in 1950 [Anon, 1999].

Coal mining began in the Burragorang Valley (at Nattai) on a small scale in the 1930s but it soon became the principal economic activity. The Nattai North, Nattai Bulli and Wollondilly collieries commenced in the 1930s and ceased operations during the early 1990s. The Valley collieries started their operation in the early 1960s and continued through until the mid 1980s. The Mt. Waratah (near 'The Crags') and Mt. Alexandra (Mittagong) collieries are located in the upper Nattai River catchment, whilst coalmines located to the north and northeast of the lower Nattai River catchment include the Brimstone Colliery and Oakdale Colliery. Coal operations in the Burragorang Valley have now ceased [Colliton, 2001]. 20 There is a broad spectrum of land uses in the catchment including urban development, agriculture and national park. Soil erosion and contaminant release were a cause of concern for the drinking water supply after a large-scale bushfire in the Nattai catchment in December 2001/January 2002 [Agnew, 2002]. During periods of high flow, the Nattai carries large volumes of fine sediments, partly as a result of land clearing in the upper catchment [McCotter, 1996]. The Sydney Catchment Authority (SCA) has two water quality monitoring stations, Crags and Causeway situated along the Nattai that measure standard parameters such as pH, DO, nutrients, thermotolerant coliforms, Chlorophyll-A and a few selected metals [SCA, 2001]. Site Crags has been found to breach the guideline range for pH, DO, total nitrogen and phosphorus and Chlorophyll-A. Thermotolerant coliforms and total nitrogen values have been found above the guideline levels at Causeway [SCA, 2001]. Nutrient concentrations in the river were high during both dry and wet weather, exceeding guidelines on most occasions. Water quality considerably improved at the inflow to the lake (Causeway), with very few dry-weather samples containing concentrations above guideline levels. All sites along the Nattai River were turbid during wet weather [SCA, 2001a]. Decline in the water quality can be attributed to the infrastructure of urban development such as STPs, Swimming Pools, Golf Courses and a Rubbish Tip [Colliton, 2001].

In the Warragamba catchment, 41% of the water flowing into Warragamba comes from the Wollondilly inflow whose catchments include Goulburn and the Southern Highlands [McClellan, 1998]. The entire catchment area is approximately 3403 square kilometers. The Wollondilly is the largest of the inflows and has the Mulwaree, Tarlo, Paddys and Wingecarribee rivers as its major tributaries. The Wollondilly catchment is characterised by broad open valleys with gentle rolling hills, which have been mostly cleared for agriculture/grazing purposes [CSIRO, 1999].

The primary hazards in these catchments derive from the impact of animal grazing with stock access to streams, the large number of unsealed roads and tracks, intensive pig and poultry enterprises, stables, saleyards, meat and wool processing [CSIRO, 2001]. The headwaters of the Mulwaree River and a tributary, Crisps Creek, also have been affected by the activities of the Woodlawn mine, which

21 produced gold, silver and zinc [Jones and Boey, 1992]. This mine was closed in 1998. It is now proposed to use the site as a waste disposal facility [AWT, 2001].

The treated effluent from Goulburn STP is pumped onto designated areas and does not go directly into the Wollondilly River. However, there are limited storage facilities for its partially treated effluent and no wet weather storage at the irrigation area. Therefore there are pronounced chances that during heavy flow the irrigated effluent, including the partially treated effluent, can be washed into the Wollondilly River [McClellan, 1998]. SCA collected samples during wet weather from the upper Wollondilly River, just downstream of Goulburn and found nutrient concentrations above recommended guidelines. At the inflow to Lake Burragorang, phosphorus concentrations were generally acceptable during wet weather, although total nitrogen concentrations were still mostly elevated. The Mulwaree River and Wingecarribee River, which flow into the Wollondilly River indicated poor water quality with pH, dissolved oxygen, turbidity, nutrient, and chlorophyll-a concentrations failing to comply with recommended guidelines [SCA, 2001].

Werri Berri (Monkey Creek) is a sub-catchment of the Lake Burragorang catchment, accounting for 2% of the total catchment area. Only 0.5% inflow come from Werri Berri, though a relatively small stream, it is of particular importance, due to its entry point close to the dam wall (approximately 4 kilometers from the offtake point for Sydney's water supply) and its fairly urbanised character. Werri Berri Creek catchment is the most developed area in the Warragamba Special Area . Forty per cent of the Werri Berri Creek catchment is developed, with the remainder retained as bushland [SCA, 1999].

Land use in the area includes unsewered residential development (principally within the towns of The Oaks and Oakdale), small rural sub-division and agriculture (predominantly livestock, vegetable growing and poultry and hobby farming). Horse Creek, which flows into Werri Berri Creek, has coalwashing activities at its headwaters. One of the larger mines in the area, the Oakdale mine, was closed in August 1999. There are still a number of mines in operation [AWT, 2001]. The impact of development in the Werri Berri Creek catchment poses a risk to the water quality of Lake Burragorang. Water quality problems have been found in the upper part of the catchment including high levels of turbidity, iron, nutrients and faecal 22 bacteria. Cryptosporidium and Giardia have been detected in storm water channels draining from the Oak township to Werri Berri Creek [SCA, 1999]. The unsewered townships could be the prime cause for such contamination [AWT, 2001]. Rural land uses such as dairying, grazing (sheep, deer, cattle and horses), market gardening, turf growing, poultry and hobby farming may also contribute to the poor water quality in Werri Berri Creek. The NSW Government has placed this area on the Priority Sewage Program [SCA, 1999].

The quality of water entering Lake Burragorang is usually of a lower quality when compared to water extracted at the offtake. The narrow shape of Lake Burragorang, combined with its large area and depth, allows a long residence time for most waters entering the lake before reaching the abstraction point at the dam wall. The long residence time allows lake processes such as sedimentation and assimilation of nutrients by living organisms to improve the water quality within the lake [SCA, 1999].

The Lake acts as a final contaminant removal area, before the water is piped to the Prospect Water filtration plant in Sydney. When the Lake fails in contaminant removal task then problems such as the pathogens Cryptosporidium and Giardia can appear in the municipal water supply, as occurred during Sydney’s “Boil Water Crisis” in 1998. Sydney’s drinking water came under scrutiny after detecting these pathogens. Water quality monitoring and assessment has increased ever since this incident took place. Water quality monitoring by SCA has focused mainly on nutrients (eutrophication) and microbiological analysis to maintain the quality safe for human health. Limited monitoring of metals (only Fe, Al and Mn) has been undertaken to assess water quality aesthetics and treatability, rather than to assess metal contaminants from an ecological health perspective. Except Al, these metals are not considered to pose an ecological or human health risk compared to other metals such as Cu, Zn, Cd and Pb [CSIRO, 2001]. The anthropogenic activities discussed in Table1.1 within different catchments pose a potential risk of metal pollution to water and sediment quality in Lake Burragorang. Increased concentrations of Mn and Fe have been reported in deeper water towards the end of stratified period the cause of which could be the depletion of oxygen in the hypolimnion and the release of metals from the sediment [SCA, 2001a]. The audit and inquiry on Sydney water [CSIRO, 2001] recommended that the investigation be 23 expanded to bottom sediments also as the cause of contamination could be the resuspension of settled material during major inflow. It is therefore very important to study the limnological processes occurring in the Lake Burragorang, to ensure the best quality water is delivered to Sydney.

The chemical characteristics of the fine sediments in the river system affect instream habitats, which influence ecological conditions. Various pollutants− particularly metals and hydrocarbons assume great significance in pollution study of aquatic systems − they can accumulate in fine river sediments and may affect the health of the stream ecosystem. The accumulation of heavy metals in the bottom sediments of water bodies and the remobilisation of these substances from the latter are two of the most important mechanisms in the regulation of pollutant concentrations in an aquatic environment [Linnik and Zubenko, 2000]. In the past, however, water quality studies focused mostly on the detection of contaminants in the water column and ignored the fact that sediments may act as large sinks or reservoirs of contamination [Horowitz, 1991; Loring, 1991; USEPA, 2000]. Many past studies also failed to recognise that remobilisation of metals from contaminated sediments can cause water quality problems [USEPA, 1999].

As part of the Warragamba catchment-monitoring scheme, number of water quality reports have been compiled by catchment authorities and local councils on inner and outer catchment of Warragamba. However, scant studies have been done on its sediment quality. Few significant studies have been carried out in recent years on Lake Burragorang subcatchments to examine the distribution and concentration of trace metals and likely sources of contamination.

A comprehensive survey conducted by Australian Water Technology [AWT, 1994] indicated that most trace metal concentrations (As, Cd, Cr, Cu, Pb, Ni and Sn) were below guidelines [ANZECC/NHMRC, 1992] for concentrations of metals in contaminated soils. Nine out of the 46 sites sampled had zinc concentrations exceeding the ANZECC criterion − mostly in the upper Coxs River catchment. The highest concentrations of reactive zinc were found in sediments from Marrangaroo Creek and Blackmans Creek. Most of the 46 sites had manganese concentrations exceeding the ANZECC criterion, again in the upper catchment of the Coxs River [Young et al., 2000]. 24 Harrison et al. [2003] investigated the core and surface sediments from Tonalli River, approximately 12 km west of the lower catchment of the Nattai River. They established the temporal variability of metal concentration through 210Pb dating and compared with historical records, rainfall and bushfire. Their study concluded that heavy wet seasons greatly influenced the sediments grain size, organic content and trace metal concentrations. Spatial distributions indicated that greater concentrations of trace metals were associated to local mining processing sites. Birch et al [2001] studied distributions of trace metals in the fluvial sediments of the Coxs River (the main northern catchment to Lake Burragorang) and observed that increase in specific trace metals could be related to anthropogenic sources, ranging from urban settlements through to Sewage Treatment Plants (STP) and local coalmines of the area. The spatial and temporal distributions of contaminated fluvial sediments within Nattai catchment were studied by Colliton [2001] to determine the impact of urban settlement and identify influential contamination sites. The study showed that there is strong correlation between the concentration of trace metals in the sediments and the geological formations of Nattai catchment. The study also indicated the relationship between major fire events and catchment erosion resulting in increase sedimentation with coarser composition. Agnew [2002] determined the effects of recent bushfires on sediment and pollution transport in the Nattai catchment. The study also examined the relationship between bushfire and sedimentary charcoal record.

Inspite of the significance of Lake Burragorang to large population of the Sydney, no systematic metal distribution and speciation study of its water and sediments had been carried out in the past. Keeping this view in mind a detailed study was undertaken during 2002-2004 to investigate the distribution of heavy metals (As, Cd, Cr, Co, Cu, Fe, Pb, Mn, Hg, Mo, Ni, Se, V and Zn) and their speciation in Lake Burragorang sediments to understand their bioavailability and toxicity to aquatic system of the lake. The bed sediment samples from various preselected sites were collected and analysed for distribution of heavy metals and their speciation. The selection of the sampling stations was based on the consideration of maximum representativeness and approachability.

25 1.3. Report Organisation

For convenience and clarity of presentation, the subject matter of the thesis has been divided into following seven chapters.

First chapter provides a brief background of environmental pollution with reference to metal pollution. It includes a detailed description of study area and Lake Burragorang including major inflow in the lake and their catchment activities related to possible source of contamination. Based on the available information, the objectives of the work embodied in the thesis have been defined.

Second chapter gives details of the mode of sampling, preservation of samples and methodology used for the analysis of physico-chemical parameters. The procedures followed for the speciation of metals in bed sediments are also given. The methodology adopted for sedimentation rate and nutrient analysis is also discussed. It also includes detailed description of instruments used for different analysis.

The relevant literature available on metal speciation using sequential extraction and results of metal distribution and their bioavailability on sediment grabs samples collected from Lake Burragorang have been discussed in third chapter. Using sequential extraction procedure given by Tessier et al [1979], the metals are differentiated into five categories, adsorptive and exchangeable, bound to carbonate phases, bound to reducible phases (iron and manganese oxides), bound to organic matter and sulphides and detrital or lattice metals. The spatial variations and remobilisation ability of various chemical forms have been discussed.

The fourth chapter presents the results of organic matter and carbonate contents of the lake sediment including depth profile of nutrients and metals. It also deals with speciation of sediment core using Simultaneously Extracted Metal (SEM) and Acid Volatile Sulphide (AVS) ratio. The method is based on the fact that when the ratio of the toxic heavy metals (SEM) to reactive sulphide (AVS) is less than 1, no toxicity is predicted for the sediment.

The fifth chapter describes the sedimentation rate results on few preselected locations of Lake Burragorang. The age of sediments was obtained using 210Pb dating method as described by Brugam [1978] and thus the variation in metals and

26 nutrients in the sediments with age was established and compared with published historical record, rainfall records and bushfire data.

The sixth and seventh chapters include conclusion and references, respectively

27

Chapter II. Materials and Methods

Lake Burragorang, impounded by Warragamba Dam, is one of the largest domestic water supply storages in the world, holding 2,057,000 million liters of water - over four times the volume of Sydney Harbour. Such a large storage is essential during the extended periods of drought that the Sydney region experiences. A record drought from 1934 to 1942 necessitated the construction of Warragamba Dam to provide a reliable water supply for Sydney’s growing population. Lake Burragorang, formed behind Warragamba Dam, has a surface area of 7,500 ha and collects water from a 905,000 ha hydrological catchment area.

2.1 Field Sampling

Sediment grabs and cores samples were collected for speciation and sedimentation study. Sixteen sampling locations (Fig 2.1) were chosen to cover the 7,500 ha lake area as well as to study the effect of inflow from surrounding rivers. Sampling locations have been discussed in more detail in the following chapters. Recommendations of Batley [1989] have been followed in this study for sample collection, handling and storage.

2.2 Sediment Grab

Bottom sediment samples were collected by Ponar Petite grab in May 2002 (Figs 2.2 and 2.3). The sediment grab was lowered through the water column at approximately 1m per second to minimise disturbance of the sediment by a “bow wave” of water in front of the grab. The sediment grab collected sediment from the lake bottom of approximately 30cm x 20cm surface area, to a maximum depth of around 10 cm.

Composite samples of the sediment were collected using a polyethylene scoop. The sediment was then placed into polyethylene plastic bags, which were then tightly sealed, labelled and placed under ice in an insulated box. Samples from sediment

28 grabs were placed in a freezer below –10 °C on arrival in the laboratory until analysed.

Fig-2.1. Locations of sediment core and grab samples in Lake Burragorang

29

Fig 2.2. Ponar Petite sediment grab sampler

Fig 2.3. Sediment grab sample collected from Lake Burragorang

30 2.3 Sediment Core

Sediment cores were collected in November 2002 and June 2003, using KB messenger-operated gravity type core sampler (Figs 2.4). The sediment corer enabled sediment cores up to 45cm in length and 4.3cm in diameter, enclosed within an acrylic inner tube and capped at either end with a polyethylene cap. After collection cores were labelled and kept upright (in order to preserve the natural stratigraphy) under ice in an insulated box. In general, duplicate cores were taken at each sampling location. On return to the laboratory the sediment core tubes were placed on the purpose built sediment core extrusion device (Fig 2.5 and 2.6), and the contents of the tubes forced out through the core stripper.

A B Fig 2.4. A) KB Sediment corer B) Sediment in an acrylic sediment core tube

31

Fig 2.5. Sediment core extrusion device

Fig 2.6. Top of sediment core stripper

32

Fig 2.7. Details of sediment core stripper

The core stripper (Fig 2.7) was designed to remove the outmost 3 or 4 mm of the sediment from the sediment core (i.e. the sediment that had been in contact with the inside of the acrylic tube). This was done to avoid the problem of smearing, which occurs when the outside part of the sediment core is smeared along the inside of the acrylic tube as the sediment is forced out of the tube. This will prevent the mixing of sediments of different ages.

Each sediment cores were sliced at 5 cm interval throughout the entire length, homogenised, and stored in separate labelled polystyrene containers below -10°C until required for analysis.

2.4 Analytical Methods

2.4.1 Moisture Content

Approximately 10 g of sediment sample was placed in a previously dried (105 °C) crucible and dried in an oven at 105 °C to constant weight. The moisture content was then calculated as follows

33 wet sedim ent weight − dry sediment weight Moisture content = x100 wet sediment weight

2.4.2 Organic matter and Carbonate content

Batches of sediment samples were heated in 2 separate sessions in the laboratory's furnace: (1) for 4 hours at 500 °C (for the removal of the sample's organic content); (2) for 2 hours at 1000 °C (for the removal of the inorganic carbonate content) [Dean, 1974]. The crucibles were allowed to cool down to approximately 100 °C in the furnace, before being subsequently conveyed to several large desiccators and allowed to cool at` room temperature prior to weighing. The porcelain crucibles were pre-heated in a furnace to 1000 °C.

Approximately 2 g of dry sediment sample were added to each of the crucibles. The difference in mass of the samples was recorded following the completion of each of the 2 heating phases.

initial crucible weight − final crucible weight % Organic matter = x 100 initial sedim ent weight

2.4.3 Total Nitrogen and Phosphorus

Total nitrogen and total phosphorus samples were analysed on a Lachat Quickchem 8000 (Lachat Instruments, USA) flow injection system. Briefly, 0.2 gram of sediment sample was digested with sulphuric acid (H2SO4), potassium sulphate o K2SO4 and copper sulphate (CuSO4.5H2O) at 390 C for 3 hours in a block digester. After cooling the sample was diluted and subjected to Flow injection analyser. During the digestion, the phosphorus in the samples is converted to orthophosphate. In the chemistry manifold the orthophosphate reacts with ammonium molybdate and antimony potassium tartrate under acidic conditions to form a complex. This complex is reduced with ascorbic acid to form a blue reduced phosphomolybdenum compound, which absorbs at 880 nm. The absorbance is proportional to the concentration of orthophosphate in the digest [Lachat Instruments, 2000].

34 Digestion process convert the nitrogen in the sample into ammonium cation, which is then injected and heated in stream of salicylate and hypochlorite to produce blue color complex, which is proportional to the ammonia concentration [Lachat Instruments, 2003].

2.4.4 Acid Extractable Metal

The fine fraction of silt/clay particles was chosen for the metal analysis as the higher concentration of heavy metals generally accumulate on smaller size (<63 μm) grain fractions [Whitney, 1975; Harding and Brown, 1978; Horowitz and Elrick, 1987; Kersten and Forstner, 1989]. Wet sediments were analysed as the drying process is known to significantly alter metal speciation [Batley, 1989; Kersten and Forstner, 1989; Jones and Turki, 1997].

A portion of each bulk sample was size-normalised by wet sieving through a 63 μm nylon mesh screen. Subsamples of homogenised wet sediment, equivalent to 1g dry weight (moisture content determined on separate aliquot) were digested with reverse aqua-regia in an ultrasonic bath at 60 °C for 45 minutes followed by hotplate treatment at 145 °C for 45 minutes [Siaka, 1998]. Blank and standard reference samples were also digested in the same way.

2.4.5 Speciation

2.4.5.1 Sequential Extraction

For sediment grabs, sequential extractions were performed to determine the amount of metals that were associated with different chemical fractions of the sediment. The procedure performed, follows the guidelines and parameters published in previous work by Tessier [1979]. This scheme consists of five successive extraction steps (Fig 2.8). Wet sediments were used, as the drying process is known to significantly alter metal speciation. All sediment samples were wet sieved through 63 μm nylon mesh screen and homogenised

Step I –Exchangeable Fraction - Wet sediments equivalent to 1 g dry weight were weighed in clean dry centrifuge tubes and shaken at room temperature with 10 mL of 1M MgCl2 at pH 7 for 1 hr. The suspension was centrifuged at 3000 rpm for 20

35 minutes and the supernatant removed for later analysis. The remaining sediment was washed with Milli-Q water before the next extraction step.

Step II- Carbonate Fraction – The sediment remained in the centrifuge tubes after step I was extracted by 10 mL of 1 M NaOAc at pH 5 for approximately five hours at room temperature. Again, the suspension was centrifuged, the supernatant saved for analysis, and the remaining sediment washed.

Step III – Fe-Mn Oxide or Reducible Fraction- The oxides of iron and manganese were targets in this step. The extraction was performed using 20 mL of

0.04 M NH2OH-HCl in 25% CH3COOH for 6 hours at 100 °C.

Step IV- Organic or Oxidisable Fraction- Organic matter was targeted in the next extraction using 5 mL of each 0.02 M HNO3 and 30% H2O2. The solution was extracted at 100 °C for 5 hours at pH 2. On cooling 3.2 M NH4OAc in 20% HNO3 was added and then shake for 30 min with continuous stirring.

Step V- Residual Fraction- In the final step the remaining residues after 4th extraction were digested with 10 mL reverse aqua regia in an ultrasonic bath at 60 °C for 45 minutes followed by hotplate treatment at 145 °C for 45 minutes. Any metal intimately associated with phases such as silicates will not be extracted since HF was not used in the residual extraction step.

The sequential leaching procedure was carried out without delay once started, and sample storage during the process (e.g. overnight) was at 4 °C. The sample handling for step I-III was performed in a glove box under nitrogen atmosphere, and all reagents were deoxygenated with oxygen-free nitrogen prior to use (Fig 2.9). The centrifuge tubes were sealed under nitrogen in the glove box prior to removal for shaking etc [Kersten and Forstner, 1986].

2.4.5.2 Simultaneously Extracted Metal (SEM) and Acid Volatile Sulphide (AVS)

SEM-AVS method was used to assess the potential toxicity of the sediment cores.

All sediment samples were wet sieved through 63 μm nylon mesh screen and

36 1g SEDIMENT

SHAKE 1Hr +1 M MgCl2 (10 ml), pH 7.0, 1Hr 25 ± 2 ºC

CENTRIFUGE

STIR 5 Hr RESIDUE + 1M NaOAC (10 ml) pH 5.0, 25 ± 2 ºC SUPERNATENT (EXCHANGABLE CENTRIFUGE FRACTION)

HEAT, 100 ºC, RESIDUE + 0.04 M NH2OH.HCl SUPERNATENT 6 hr IN 25% CH3COOH (20 ml) (CARBONATE FRACTION) CENTRIFUGE

RESIDUE + 0.02 M HNO3 (5ml) + 30% H2O2 (5 ml), pH 2.0, 100 ºC, 2 hr; 5 ml SUPERNATENT 30% H2O2 pH 2.0, 100 ºC, 5 hr; 3.2 M (Fe-Mn OXIDE NH4OAC IN 20% (v/v) HNO3, FRACTION) CONTINUOUS STIRRING, 30 min

CENTRIFUGE

DIGESTION RESIDUE + 10 ml REVERSE AQUA REGIA, 45 min, 60 °C on ULTRASONIC BATH, THEN on HOT PLATE for 45 min at 145 °C

RESIDUAL FRACTION SUPERNATENT (ORGANIC FRACTION)

Fig 2.8. Flow chart of sequential extraction scheme for sediments metal speciation

37

Fig 2.9. Extruding a sediment core in a glove box under nitrogen

homogenised. A rapid screening method [Simpson, 2001] was used to determine acid volatile sulphide in sediments.

For AVS method, in a nitrogen gas filled glove box, 0.1 g sample of sediment was accurately weighed, and transferred to a centrifuge tube. 50 mL of deoxygenated Milli-Q was added, followed by 5 mL of methylene blue reagent (MBR was prepared by first dissolving 2.8 g of N-N-dimethyl-p-phenylene-diamine hemioxalate salt in 1000 mL of cold sulphuric acid solution (670 mL H2SO4, 330 mL Milli-Q). This solution was then mixed with 200 mL of 0.020 M acidic ferric chloride solution (5.4 g FeC13.6H2O dissolved in 100 mL HCl and 100 mL Milli-Q. The final MBR solution was approximately 22 N and was stored in an amber bottle (stable for at least one month) and the centrifuge tube was capped and inverted few times to mix. After 5 min the sample was centrifuged (2 min, 2,500 rpm) and then allowed to sit for 90 min for the methylene blue colour development. The centrifugation and colour development stage was performed outside the nitrogen gas-filled glove box with the centrifuge caps tightly sealed. During this period, care

38 was taken not to significantly disturb the sediment (i.e., no further shaking) because MBR adsorbs to sediment particles. After colour development (90 min), standards and samples were analysed at 670nm with an ultraviolet-visible spectrophotometer.

Simultaneously extracted metals (SEM) were extracted in 1M HCl [DiToro et al., 1992] for 30 min at room temperature. When the ratio of the toxic heavy metals (SEM) to reactive sulphide (AVS) is less than 1, no toxicity is predicted for the sediment. Metal concentrations in all solutions were determined using ICP-AES and ICP-MS.

2.4.6 Sedimentation Study

Core chronologies using 210Pb analysis was first suggested in 1963 by Goldberg [1963] and was first applied to lake sediments by Krishnaswamy et al. [1971]. The total 210Pb activity was determined by measuring its granddaughter 210Po, which was assumed to be in secular equilibrium with 210 Pb. Supported 210Pb was approximated by measuring 226 Ra activity.

Approximately 2g of each sample (dry weight) were spiked with 209Po and 133Ba yield tracers to determine the chemical recovery of 210Po and 226Ra respectively. The samples were leached with hot acid and refluxed for 12 hours to remove organic matter as it interferes with the analysis. Ether extraction was then performed to remove excess iron from the sample. The resulting aqueous fraction was evaporated to dryness to concentrate the radionuclides. Polonium was auto deposited onto silver discs while radium and barium were precipitated as colloidal precipitate and collected on a 0.1μm filter.

Once separated and concentrated onto a source, the radioactivity content of these radioisotopes was measured. The polonium and radium sources activity was measured using alpha spectrometry (ORTEC alpha-spectro meter). The radium source provides a measurement of the supported 210Pb activity whereas 210Po activity is in equilibrium with total 210Pb activity. Calculation of the unsupported 210Pb values was carried out by subtraction of the supported 210Pb activity from the total 210Pb activity. The sedimentation rates were calculated using the modified constant initial concentration (CIC) model method described by Brugam [1978] and using the formula: 39 1 ⎛ Ao ⎞ tI = * In ⎜ ⎟ y ⎝ At ⎠ where, Ao = unsupported 210Po at the sediment surface in decays per minute per gram (dpmg-1) of dry sediment; At = unsupported 210Po activity at time t in dpmg-1 of dry sediment; y = decay constant of 210Po (0.03114) in year -l tI = difference in age between surface sediment and sediment at depth in years.

This equation is applied to sections of the core under the assumption that within each section, the flux of unsupported 210Po was constant.

2.4.7 Statistical Treatment of Data

The different results reported in the thesis are the average of minimum of two determinations. Blank determinations were carried out wherever necessary and the corrections were made if required. During the analysis for different parameters blanks, duplicates, spikes and standards were processed on 5% basis. The percentage recovery for spiked samples in metal determinations ranged from 94 to 104%, which indicate that the results are accurate and unbiased. Relative percent difference of duplicate measurements was less than 10%, which is a satisfactory precision.

The uncertainty associated with various analysis (organic matter, carbonate content, nutrients and metals) was performed by calculating standard deviation and coefficient of variation (CV) on randomly selected samples. The results are shown in Table A1. Satisfactory precision were consider as CV values for all variables were <10%. The analytical procedure for the determination of acid extractable metal concentrations was checked by means of analysis of standard reference samples- AGAL-10 (reference sediments from Hawkesbury River, NSW) and AGAL-12 (biosoil, a mixture of soil and dried sewage sludge). These reference samples were obtained from the Australian Government Analytical Laboratories (Pymble, NSW). The data obtained from the analysis of the reference materials is reported in Table

40 2.1. The observed values obtained were within, or close to certified values. The percentage recovery for all metals ranged between 75% and 107%.

Systematic errors associated with radioisotope counting were directly calculated by the computer interfaced to the mass spectrometer and were incorporated into the errors quoted with the activity result sheets.

Table 2.1. Comparison of reference material values with obtained results

Metal AGAL-10 AGAL-12 (mg/kg) Observed Certified Observed Certified values values values values As 16.5 ± 0.7* 18.7 ± 0.8 2.98±0.2 3.54 ± 0.4 Cd 8.06 ± 0.3 9.55 ± 0.65 0.71± 0.2 0.77± 0.4 Co 7.87 ±0.3 9.3 ± 0.8 7.2 ± 0.5 8.61± 0.8 Cr 67.8 ±3.8 85.62 ± 12.2 28.2 ± 4.8 33 ± 2.2 Cu 18.8 ±2.7 22.55 ±1.6 113.7 ± 1.5 150 ± 2.6 Fe 19106 ±422 20163 ± 2356 27086 ± 126 25206 ± 1500 Hg 10.04 ±0.4 11.77 ± 0.2 0.41±0.1 0.53 ± 0.5 Mn 188 ±5.2 247 ± 9.3 398±8.1 497 ± 24 Mo 7.6 ±0.7 9.37 ± 1.4 1.15±0.6 1.53 ± 0.4 Ni 17 ±4.1 18.2 ± 3 16.1±0.9 17.2 ± 1.2 Pb 33.3 ±2.3 39 ± 5.2 24.7±4.1 31.4 ± 1.6 Se 11.3 ±0.35 11.67± 0.7 1.22±0.1 1.56 ± 0.3 V 22.2 ±1.2 27.1 ± 0.8 25.1±0.8 31.8 ± 1.6 Zn 52 ±5.4 55.1 ± 3 157.4±3.6 182 ± 7.1

n =5 *= Standard deviation

41

Chapter III. Distribution of metals and speciation in sediment of lake Burragorang using sequential extraction

3.1 Introduction

Heavy metal pollution of aquatic systems is a serious problem and has attracted a lot of attention of scientific community worldwide. Unlike the organic pollutants, heavy metals are not removed by natural processes of decomposition, on the contrary, they may be enriched by organisms (biomagnification) and can be converted to organic complexes, which may be more toxic. It has been widely recognised that identification of metal forms or species is necessary to understand their bioavailability and toxicity in the system [Fytianos, 2004; Korfali, 2004; Rauret, 1988; Li, 2000]. The total metal will only be able to provide information about the pollution if the background level or geochemical composition is known; metal origin (natural or anthropogenic) is rather difficult to predict. Thus to assess the environmental impact of sediments the determination of trace metal is not sufficient in itself [Salomons and Forstner, 1980]. The chemical form of the metal in the sediment ultimately determines the behaviour and mobilisation ability of the metal in the environment.

The concentration of metals in any particular sediment will depend upon many interacting factors such as, sources of sedimentary materials, the processes, which lead to the presence of suspended metal containing particles in the water column and the hydraulic and chemical factors [Gadh et al., 1993]. When a trace metal entered into riverine system its distribution among various compartments may be due to variety of processes including solubilisation, competitive chelation, precipitation, sedimentation, adsorption and uptake by planktonic living organisms [Kramer, 1991].

Metals in the sediments are mainly associated with detrital, authigenic and biogenic components. Aluminosilicate minerals ultimately derived from the rocks by

42 weathering and supplied to lakes and oceans by rivers, ice and on-shore sediments are mainly detrital. Biogenic sediments may contain calcareous and siliceous skeletal matter and finely dispersed organic matter. Authigenic component consists of ferromanganese oxides, precipitated carbonates and sulphides and interstitial water. Precipitated hydrous manganese and iron oxides are abundant in all the oceans of the world, in shallow marine environments and in many temperate lakes [Cronan, 1976]. Ferromanganese precipitates are usually enriched in trace metals compared with detrital sediments but the degree of enrichment varies according to the depositional environment and the particular trace metal. Inorganic precipitation of carbonates is believed to exert some control over trace metal levels in the water column [Calvert, 1976]. Biological processes within the deposited sediments are mainly responsible for authigenic sulphides. Decomposition of organic carbon and sulfate ions leads to the formation of hydrogen sulphide. Iron and other metal cations may then be precipitated to a degree, which depends on the sulphide ion concentration and the strength of the competing bonds to organic complexes [Timperley and Allan, 1974; Calvert, 1976; Jackson, 1978]. Since sulphides are invariably produced in organic rich reducing environment where the organic matter and sulphide are intimately mixed, it is difficult to determine the partitioning of heavy metals between these two sediment components. Sediment interstitial water, or pore water, is defined as the water occupying the spaces between sediment particles. Interstitial water differs in composition from the overlying water. Trace metals are generally enriched in interstitial waters [Elderfield and Hepworth, 1975].

This information on sediment characteristics helped researcher to develop the leaching scheme for partitioning of metals among various forms in which they might exist in sediments. In the literature, numerous sequential extraction schemes are described [Tessier et al., 1979; Sposito et al., 1982; Welte et al., 1983; Clevenger, 1990; Ure et al., 1993.; Campanella et al., 1995; Howard and Vandenbrink, 1999] to study the mobility and availability of the metals in the sediments. The sequential extraction procedure developed by Tessier et al. [1979] is one of the most thoroughly researched, which furnishes detailed information about the fractionation of trace metals and widely used procedures to evaluate the possible chemical associations of metals in sediments and soils [Li et al., 2000]. International Union Of Pure And Applied Chemistry (IUPAC) technical report also recommend the method of

43 sequential chemical extraction as the least sophisticated and most convenient technique available for a speciation assessment [Hlavay et al., 2004].

However, It is important to understand what is happening during extraction to minimise the risk of producing artifacts and choose standard procedures to ensure that results are comparable.

The mechanism of accumulation of heavy metals in the sediment components may lead to the existence of metals in the following broad categories [Gunn et al., 1988].

- Pore water

- Adsorptive and exchangeable

- Bound to carbonate phases

- Bound to reducible phases (iron and manganese oxide)

- Bound to organic matter and sulphides

- Detrital or lattice metals.

The geochemical behaviour of trace metals and their chemical forms can be ascertained with the help of fractionation. Assuming that bioavailability is related to solubility, then metal bioavailability decreases in the order: exchangeable > carbonate > Fe–Mn oxide > organic > residual [Tessier et al., 1979; Ma and Rao, 1997]. The fractions introduced due to human activities include the adsorptive and exchangeable and bound to carbonates which are considered to be weakly bound and may equilibrate with aqueous phase thus becoming more rapidly bioavailable [Gambrell et al., 1976; Gibbs, 1977; Young and Harvey, 1992]. On the other hand, the metal present in the inert fraction, being of detrital and lattice origin, can be taken as a measure of contribution by natural sources [Salomons and Forstner, 1980] which are not easily mobilised. The Fe-Mn oxide and the organic matter have a scavenging effect and may provide a sink for heavy metals. The release of the metals from this matrix will most likely be affected by the redox potential and pH [Gambrell et al., 1976].

During the past 20 years sequential extraction schemes have been employed by several researchers for the determination of binding forms of trace metals in different sediments of the various rivers. 44 Li et al [2000] studied the chemical forms of four heavy metals (Zn, Cu, Ni and Co) and their spatial distribution using sequential extraction in the sediments of the Pearl River Estuary, China. The sequential extraction results showed that Zn, Ni and Co in the top sediments were mainly associated with the residual and Fe–Mn oxide fractions whereas the major geochemical phases for Cu were the organic and residual fractions.

Fytianos and Lourantou [2004] applied the sequential extraction procedure for the determination of the distribution of seven elements (Cd, Pb, Cr, Cu, Mn, Zn, Fe) in sediment samples collected from two lakes, Volvi and Koronia, located in North Greece. Based on the results obtained at one sampling point in lake Koronia and two sampling points along the lake Volvi, authors have concluded that the water of the two lakes is not polluted. There were no significant changes in the individual seasonal concentrations of elements in this monitoring period. Cd, Pb, Cu and Cr are associated with the oxidisable, carbonates and residual fractions. Zn and Fe are associated with residual and reducible fractions. The metals most easily extracted in the samples analysed in both lakes are Pb, Cr, Cd, Cu and also Mn in the case of Koronia Lake.

Korfali and Davies [2004] analysed speciation of metals in sediment and water in one of Lebanon’s river the Nahr-Ibrahim, whose basin is underlain by limestone and its water is dominated by carbonate species due to the high pH and alkalinity values. Sequential chemical fractionation scheme was applied to the -75 mm sieved sediment fraction. The data showed that the highest percentage of total metal content in sediment is for Fe in the residual fraction followed by moderately reducible fraction, Zn and Pb in the carbonate and in the moderately reducible fractions and Cd primarily in the carbonate fraction. Jones and Turki [1997] studied the distribution and speciation of heavy metals in surficial sediments from the tees estuary, England. Cr, Pb and Zn are associated with the reducible, residual, and oxidisable fractions. Cu is associated with the oxidisable and residual fractions, and Co and Ni, which are not highly enriched, are hosted mainly by the residual phase.

Akcay et al. [2003] investigated heavy metal pollution and speciation in the sediments of two economically important rivers of Turkey, Gediz and Buyuk 45 Menderes (BM). Pb enrichment in Gediz River sediments has an exchangeable character and represents potential pollution in this river. As in the Ni speciation study, this metal was found bound to silicates. Thus, it was concluded that both rivers have no anthropogenic source of Ni pollution. The Cu contents of Gediz River were higher than the Cu content of BM River, especially in Kemalpasa-Manisa region and this is a potential pollution risk for this region. Speciation studies prove that the industrial wastes may cause this pollution. Leaching, extraction and ion- exchange studies show that Mn compounds, which are pollution indicators, occur primarily in the first three fractions in Gediz River. It is suggestive that bioavailability of Mn in organic matter of sediments is lower; on the other hand exchangeable manganese species are abundant especially in the Gediz river sediments. These results show that Mn pollution might have originated from a kind of pesticide, which contains Mn and is used widely in this region. Cr analysis indicated the pollution in Gediz River. High Cr (VI) values confirmed that the pollution originated from industrial activities is crucial. However, in the BM River sediments Cr species are located mainly in the fourth and fifth fractions, which may originate from the geochemical composition of this region. The speciation data for Co suggests a weak pollution risk in both rivers.

Speciation of Pb, Zn, Cr, Co, Ni, Cu have been determined in the sediment of river Jhanji, India by Baruah et al. [1996]. Their results showed the significant association with residual fraction. Fe-Mn oxide fractions also scavenge a good portion of metals in them. They have not reported any significant association with organic fraction except copper.

Kwon and Lee [2001] studied the ecological risk assessment of sediment in wastewater discharging area at Masan Bay, South Korea by means of metal speciation. In this study exchangeable fraction of superficial sediment (0–2 cm layer) was detected with Zn 35.09%, Pb 5.30%, Cu 0.86%, Cr 0.01% and Fe 0%. However, exchangeable fraction of deep layer sediment (15–20 cm) was not observed for all metals analysed. Deeper sediments were found to have more residual fraction and bioavailable phases decreased with depth, which indicate the seriousness of wastewater discharge effect in this enclosed bay. Stone and Droppo [1996] analysed distribution of Pb, Cu and Zn in the size fractioned riverbed sediments in two agricultural catchments of southern Ontario, Canada. The major 46 accumulative phases for Pb, Cu and Zn were carbonates, Fe-Mn oxides and organic matter but relative importance of each phase varied for individual metals and grain size. The extraction data show increasing bioavailability of metals with decreasing grain size.

Chemical forms of cadmium, copper, lead and zinc have been determined in the bed sediments of River Yamuna by Gadh et al [1993]. Sediment characteristics do not show any significant variation except that carbonate content is consistently higher in the post-monsoon season. The speciation profiles for a particular metal show a similar trend throughout the stretch with no significant spatial variation. Cadmium is mostly associated with carbonate content and thus has a possibility of becoming readily bioavailable. Major fraction of copper is bound to organic matter while that of zinc to Fe-Mn oxide. Thus they cannot be easily leached out and pose less environmental risk. Major percentage of lead is found in the Fe-Mn oxide fraction, moderate contribution being made by carbonate and residual fractions. The total lead in the sediments is higher, therefore even a small fraction of lead bound to carbonate content can pose problems to the ecosystem. There are good correlations between the different constituents and the major metal fractions associated with it.

As already discussed in Chapter I very few references are available on Lake Burragorang sediments. Some studies have been done on its tributaries, which concentrate on metal analysis [AWT, 1994; Birch et al., 2001; Colliton, 2001; Agnew, 2002; Harrison et al., 2003]. Siaka [1998] investigated Coxs River catchment sediments for speciation of trace heavy metals using a four step sequential extraction procedure [McConcie, 1995].

Though Lake Burragorang is very important lake yet no study has been carried out on the distribution and speciation in the sediments of Lake Burragorang. Even Sydney Catchment Authority has not undertaken any monitoring of sediments in the catchment for chemical contaminants [CSIRO, 2002].

In the light of the importance of metal speciation, it is vital to find the species of metals in the sediments collected from the sites of Lake Burragorang. This will help to understand their bioavaialibility and toxicity to aquatic environment.

47 3.2 Study Area

The sampling locations in Lake Burragorang ranged from close to the Dam wall (DWA2), SW along the main canyon, DWA39 down the Wollondilly River and up the Coxs River DWA 18. A complete list of locations visited can be seen in Table 4.1. Depths of water ranged from 2 m to 90 m. In May 2002 a total of 11 sediment grab samples were collected from various parts of Lake (Fig 2.1). The selection of the sampling stations was based on the consideration of maximum representativeness and approachability. Sampling locations in the Lake were generally chosen to be at the same locations where routine water quality sampling had been carried out for some years (by Australian water technology on behalf of Sydney Catchment Authority) at so called “DWA” locations (Burragorang). This enabled any available historical data to be compared with that found during the sampling for this research. However, where necessary, other non-DWA sampling locations were used within the lake and termed “ UWS”. Fourteen metals were studied for their concentration and the chemical forms in which they occur.

The experimental procedures employed for the current study have been discussed in Chapter II.

3.3 Results and Discussion

3.3.1 Metal Distribution

The concentrations of As, Cd, Cr, Co, Cu, Fe, Pb, Mn, Hg, Mo, Ni, Se, V and Zn were analysed in sediment grab samples and are tabulated in Table 3.2. Arsenic, Cd, Cr, Cu, Hg, Ni, Pb and Se were selected as these metals are of major interest in bioavailability studies listed by U.S. Environmental Protection Agency (USEPA). Other metals were selected because of their potential for human exposure and increased health risk. Selection of metals is also based on the past and present catchment activities. The major pollution sources identified during the catchment audit process by SCA are extensive agriculture, mining, sewage systems, transport related, chemical, ceramics and other industries. The sources are already discussed in detail in Chapter I. It is difficult to make an overall assessment of the degree of metal contamination in estuarine and marine sediments [Rubio et al., 2000]. This is a consequence of variations in analytical procedures among studies and the presence 48 of an unknown natural background in the sediments. In the present study, two approaches were employed to evaluate the sediment pollution; comparison with the background value and sediment quality guidelines. The background values of the different elements were defined, depending on the international standards [Jones and Turki, 1997; Siaka, 1998; Johnston et al., 2002; Barciela-Alonso et al., 2003; Pazos- Capeáns et al., 2004; McCready et al., 2006; Nasr et al., 2006] and the background values estimated in this study in Chapter IV. The guidelines given by Long et al [1995] have been used to characterise contamination in sediments (Table 3.1). These researchers reviewed field and laboratory studies and identified nine metals that were observed to have ecological or biological effects on organisms. They defined ERL (effects range-low) values as the lowest concentration of a metal that produced adverse effects in 10% of the data reviewed. Similarly, the ERM (effects range- median) designates the level at which half of the studies reported harmful effects.

Table 3.1. Sediment quality guidelines for metals [Long et al., 1995]

Metal contaminants in sediments Metal (mg/kg) ERL ERM As 8.2 70 Cd 1.2 9.6 Cr 81 370 Cu 34 270 Fe* 20000 40000 Hg 0.15 0.71 Mn* 460 1100 Ni 21 52 Pb 47 220 Zn 150 410

*Screening Level Guidelines by Ontario Ministry of the Environment [Persaud et al., 1993]

Metal concentrations below the ERL value are not expected to elicit adverse effects, while levels above the ERM value are likely to be very toxic. A station is rated “good” if the concentrations of all nine metals are below the ERL limit. An

49 “intermediate” rating applies if any metal exceeds an ERL limit, and a “poor” rating signifies exceedance of an ERM limit for any metal [USEPA, 2002]. Interim sediment quality guidelines (ISQGs) have recently been introduced in Australia, which incorporate guidelines for fresh and marine water quality [ANZECC, 2000]. Effects range-low (ERL) and effects range-median (ERM) guidelines [Long et al., 1995] were re-named ISQG-Low and ISQG-High guidelines, respectively [McCready et al., 2006].

Table 3.2. Metal distribution in the Lake Burragorang sediment grab samples according to sampling points

S.No. Station As Cd Co Cr Cu Fe Hg Mn Mo Ni Pb Se V Zn mg/kg 1 DWA3 7 0.2 11 30 18 36000 <0.1 2970 0.67 19 19.6 0.5 38.6 64.4 2 DWA2 8.8 0.3 16 44 33 51600 <0.1 1530 1.1 29 33.0 0.4 59.4 107.8 3 DWA9 8.5 0.2 13 29 22 41300 <0.1 3740 0.63 22 21.4 <0.1 38.0 69.2 4 DWA12 6.4 0.2 11 21 21 26072 <0.1 2042 0.3 17 17.6 <0.1 27.0 70.0 5 DWA18 4.8 0.2 10 25 22 31000 <0.1 560 0.3 17 15.7 <0.1 29.0 68.2 6 DWA19 3.9 0.2 5.8 17 17 22400 <0.1 130 0.28 12 17.2 <0.1 18.0 60.4 7 DWA27 10 0.2 17 40 29 53300 <0.1 3050 0.3 27 29.4 <0.1 50.0 95.4 8 DWA35 5.7 0.1 10 27 16 33000 <0.1 530 0.13 16 18 <0.1 33.0 60.9 9 DWA39 3.6 0.2 11 29 18 30000 <0.1 380 0.1 18 17.1 <0.1 37.0 67.4 10 M3 6.1 0.2 15 27 34 46800 <0.1 340 0.39 19 24.0 <0.1 35.0 96.3 11 M7 3.4 0.2 14 20 22 32000 <0.1 210 0.24 17 19.6 <0.1 23.0 106.4 Others 3 1 13 30 20 40000 0.01-0.24 790 38 20 60 70 Background Lake Value Burragorang 4.7 0.2 12 23 20 28500 <0.1 660 0.25 19.7 22 0.13 37 68

Concentration of Hg and Se at all locations (except at DWA3 and DWA2) were found below the detection limit (0.1mg/kg). The highest level of metals among different locations was observed at DWA27 and DWA2 (Fig 3.1.). The metal concentration generally decreases in the order Fe >Mn >Zn >V >Cr >Pb ≅Ni ≅Cu>Co >As> Mo>Se> Cd as was reported by Fytianos and Lourantou [2004]. The Cd concentration throughout the lake was observed constant and was well below the background level [Jones and Turki, 1997; Siaka, 1998] except at DWA2. Se was detected only at DWA3 and DWA2 and its concentration was higher than background levels. Sites DWA2, DWA9 and DWA27 appeared to be most contaminated sites as almost all metal levels are above the estimated background values, however, DWA19 was found to be least polluted.

50 20

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0 Concentration (mg/Kg) Concentration

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50000 40000 30000

4000 Fe 3200 Mn 2400 1600

Concentration (mg/Kg) Concentration 800 0 DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7 sites

Fig. 3.1. The concentration of metals in the sediment grabs from Lake Burragorang

Arsenic, Cu, Mo and Zn exceeded the background limits at sites DWA18 and M3. Near Werri Berri at M7 Co, Cu and Zn concentrations were discovered higher than background limits. Samples at sites DWA 3 and DWA12 contained Mn about 3-5 times the background. Chromium and Fe concentrations were found to be higher than background in the whole stretch except at sites DWA12 and DWA19. The highest 51 concentrations of Cr and Fe (2 times of background) were found near the dam wall at DWA2 and DWA27, respectively and lowest down the Coxs arm. Overall metal distribution picture depicted that locations close to damwall and middle of the lake are more polluted compared to others. This may be attributed to proximity of sources. Werri Berri (Monkey Creek) catchment is close to the dam wall (approximately 4 km from the offtake point for Sydney's water supply), fairly urbanised and the most developed area in the Warragamba Special Area. Water quality problems have been found in the upper part of the catchment including high levels of turbidity, iron, nutrients and faecal bacteria. Cryptosporidium and Giardia have been detected in storm water channels draining from the Oak township to Werri Berri Creek. Oakdale colliery, which ceased operation in 1999, is in high-risk categories [DEC, 2005] located near the identified polluted sites in this study.

Based on guidelines given by Long et al [1995] the concentration of Cd, Cr, Hg, Pb and Zn were found below the ERL whereas Cu levels were close to ERL at M3 and DWA2.Arsenic and nickel were present at higher concentration than ERL at DWA2 and DWA9. Ni also exceeded the ERL at DWA27. Mn exceeded ERL at DWA35 and DWA18 and ERM at DWA3, DWA2, DWA9, DWA12 and DWA35. Interestingly Fe found to be above ERL at all sites and it is matter of great concern that it even exceeded the ERM at DWA2, DWA9, DWA27 and M3 which make these stations poor on rating.

3.3.2 Metal Speciation

Sequential extraction results can provide information on possible chemical forms of heavy metals in sediments. The trace metal distribution in different fractions in the sediment of the Lake Burragorang is outlined in Table3.3 and presented graphically in Figs 3.2, 3.3, 3.4 and 3.5. Mercury and Se were not considered for speciation due to their low concentration observed in lake sediments. The fractionation profiles indicate that arsenic is mostly bound within inert phase and the rest being present in the Fe-Mn oxide fraction. No significant spatial variations are observed in the speciation trends. The speciation scheme of Cd shows its association in all fractions except inert, however, at DWA35 it is completely in the exchangeable fraction. The oxidisable fraction is significant upstream, accounting for 25% of total Cd at DWA3. . 52 Table-3.3. Percentage of total metal content among the different sediment chemical fractions determined by sequential extractions

% Sites Fraction As Cd Co Cr Cu Fe Mn Mo Ni Pb V Zn 1 nd nd 6.4 nd nd 1.9 47.5 1.5 nd 0.2 nd 0.4 2 nd 40.0 24.5 1.3 1.2 7.1 32.0 1.5 13.9 6.2 0.5 16.6 DWA3 3 5.7 35.0 19.1 nd 0.7 12.0 12.5 nd 14.2 7.3 12.9 18.2 4 nd 25.0 12.7 12.7 45.2 4.3 5.1 89.6 17.8 15.0 2.4 18.0 5 94.3 nd 37.3 86.1 52.8 74.7 3.0 7.5 54.1 71.2 84.2 46.8 1 nd nd 13.1 nd nd 2.5 66.8 3.6 nd 0.5 0.0 4.0 2 nd 63.3 12.5 1.4 1.6 3.0 15.0 2.7 11.0 6.9 0.4 13.6 DWA2 3 9.1 36.7 19.4 nd 0.6 8.4 7.8 nd 15.5 8.5 18.5 21.3 4 nd nd 12.5 19.4 46.9 5.3 3.9 70.9 18.9 15.0 6.2 16.6 5 90.9 nd 42.5 79.2 50.9 80.8 6.5 22.7 54.7 69.1 74.9 44.5 1 nd nd 7.7 nd nd 1.5 49.5 nd nd 0.1 nd 0.8 2 nd 45.0 20.0 0.9 1.0 4.4 28.2 nd 14.8 4.1 0.3 15.5 DWA9 3 7.1 30.0 22.3 nd 0.4 10.4 15.5 nd 17.1 6.6 14.2 19.0 4 nd 25.0 16.2 18.6 50.0 7.6 5.7 73.0 21.3 12.3 3.0 19.3 5 92.9 nd 33.8 80.5 48.5 76.2 1.1 27.0 46.8 76.9 82.5 45.3 1 0.0 nd 11.8 nd nd 3.4 53.4 10.0 nd 0.3 nd 2.9 2 nd 55.0 20.0 1.9 1.2 7.0 27.9 10.0 14.8 6.3 0.8 14.4 DWA12 3 6.3 45.0 20.0 nd 0.7 14.5 11.8 nd 17.1 9.3 16.7 20.9 4 0.2 nd 11.8 19.4 54.7 7.5 3.1 10.0 21.3 13.1 7.0 15.9 5 93.8 nd 36.4 78.8 43.4 67.6 3.9 70.0 46.8 71.0 75.6 45.9 1 nd 55.0 20.0 nd 0.9 3.6 59.8 nd nd 1.4 nd 10.0 2 nd 25.0 13.0 1.8 1.7 4.2 16.4 13.3 10.2 7.0 0.3 11.5 DWA18 3 6.3 20.0 18.0 nd 0.7 12.4 8.9 nd 15.0 8.7 16.6 16.6 4 nd nd 11.0 21.9 55.3 6.1 4.3 nd 17.1 11.2 8.0 14.1 5 93.8 nd 38.0 76.3 41.5 73.6 10.5 86.7 57.7 71.7 75.0 47.7 1 nd 50.0 17.2 nd 1.6 2.4 56.2 10.7 nd 2.6 1.0 14.5 2 nd 25.0 10.3 1.3 0.5 2.8 13.1 nd 11.1 9.2 nd 14.6 DWA19 3 5.1 25.0 24.1 nd 1.2 12.2 10.8 nd 23.3 11.5 13.2 23.0 4 nd nd 17.2 41.1 64.0 5.9 4.1 nd 26.3 17.8 7.1 17.5 5 94.9 nd 31.0 57.6 32.6 76.7 15.9 89.3 39.4 58.8 78.8 30.4 1 nd 65.0 20.0 nd 1.4 0.9 73.9 3.3 nd 0.8 0.7 7.1 2 nd 35.0 14.7 nd 4.3 2.9 15.3 16.7 15.4 6.8 0.2 12.9 DWA27 3 5.0 nd 21.2 nd 1.0 11.8 6.9 nd 18.5 13.4 22.9 19.8 4 nd nd 10.6 22.3 49.3 2.9 1.7 nd 16.3 14.7 2.8 14.5 5 95.0 nd 33.5 77.7 44.0 81.5 2.2 80.0 49.9 64.3 73.5 45.6 1 nd 100.0 23.0 nd 1.5 3.8 68.9 7.7 nd 1.6 0.6 10.7 2 nd nd 12.0 0.4 4.8 4.4 13.6 15.4 11.0 7.4 0.3 12.6 DWA35 3 10.5 nd 17.0 nd 0.5 13.5 7.7 nd 16.0 12.2 23.0 16.7 4 nd nd 10.0 20.7 44.1 3.5 2.5 nd 15.1 16.5 2.8 11.7 5 89.5 nd 38.0 78.9 49.1 74.8 7.4 76.9 57.9 62.3 73.3 48.2 1 nd 50.0 14.5 nd 0.6 3.7 56.1 nd nd 1.3 0.9 7.6 2 nd 25.0 11.8 nd 3.9 4.4 14.7 10.0 11.3 8.6 0.5 15.5 DWA39 3 13.9 25.0 23.6 nd 0.7 13.1 12.1 nd 21.4 9.8 24.6 19.5 4 nd nd 12.7 28.1 46.5 5.5 4.2 nd 18.8 19.2 9.8 14.7 5 86.1 nd 37.3 71.9 48.2 73.2 12.9 90.0 48.4 61.1 64.2 42.8 1 nd 50.0 12.7 0.0 1.0 2.4 53.5 10.3 nd 0.8 1.8 8.1 2 nd 25.0 10.7 2.1 1.3 6.7 12.4 nd 11.4 7.2 0.3 13.9 M3 3 19.7 25.0 37.3 nd 1.3 19.7 21.8 nd 30.9 26.1 17.9 37.5 4 nd nd 13.3 26.4 48.2 3.7 3.5 nd 21.5 10.2 7.2 12.6 5 80.3 nd 26.0 71.5 48.2 67.5 8.8 89.7 36.2 55.6 72.9 27.9 1 nd 45.0 17.9 nd 1.9 3.3 61.4 12.6 nd 3.2 4.8 17.0 2 nd 25.0 10.0 0.7 4.3 9.1 8.6 8.4 13.3 9.4 nd 14.2 M7 3 14.7 20.0 37.1 nd 1.7 54.4 13.3 nd 33.4 25.2 17.9 36.1 4 nd 10.0 14.3 39.4 51.8 9.4 4.0 nd 25.6 11.8 6.5 12.6 5 85.3 nd 20.7 60.0 40.2 23.8 12.6 79.1 27.8 50.3 70.8 20.0

Note: 1 -Adsorptive and exchangeable, 2-Bound to carbonates, 3-Bound to Fe-Mn Oxides, 4-Bound to Organic Matter, 5-Residual or Detrital nd- None detected (below detection limits 53 and DWA9, but decreases towards the Werri Berri Creek at M7, where it accounts for only 14% of the total Other than that Cd is dominated by first three fractions. Cobalt present in all fractions, chiefly residual (42.5-20.7%) and Fe-Mn oxides (37.1-17.0%) fraction dominates. The results are in accordance with those reported by others [Baruah et al., 1996; Jones and Turki, 1997; Li et al., 2000]. The reducible fraction increases at sites M3 and M7, contained 37% of total Co and first three fractions (≥ 60%) dominate over organic and residual which means the high pollution risk around this location [Akcay et al., 2003]. The dominant fraction of Cr is residual accounting for 57-86% of the total Cr. The present study indicates Cr association with oxidisable fraction as well, which is previously reported [Davidson et al., 1994; Galvez-Cloutier and Dube, 1998; Takarina et al., 2004]. It is probably present in chromites and heavy minerals [Sager, 1992]. Thus, chromium in the residual fraction is buried in the bottom sediments as insoluble compounds and cannot reenter circulation. Copper also appeared in the same pattern as Cr, being dominated by residual and oxidisable fractions. A low percentage is also found in the exchangeable, carbonate and reducible phase at few sites. The dominant association with residual fraction can be correlated with the study of Tessier [1979] in St.Marcel and Pierrevilla sediments and Gibbs [1977] on Amazon and Yukon River sediments. Cu’s association with organic matter is probably due to its high complexing tendency for organic matter. The observed pattern with Cu is similar to those found by Salomons and Forstner [1980]; Tessier et al. [1980]; Rauret et al. [1988]; Jardo and Hickless [1989]; Pardo et al. [1990] and Gadh et al. [1993].

However, Baruah et al. [1996] and Jha et al. [1990] reported that copper in sediments of Jhanji River at Assam and Yamuna at Delhi, respectively shows preference for the Fe-Mn oxide fraction. Chemical discrimination between these two phases is difficult [Kersten and Forstner, 1989] but the affinity of Cu for organic particles and coatings is well known, sewage for example scavenging Cu strongly from seawater [Comber and Gunn, 1995]. Thus, it is in the organically bound form that Cu is most likely deposited at the sediment surface.

Iron is the most abundant metal in all sediments because it is one of the most common elements in the Earth’s Crust. Speciation of Fe depends upon source and its

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Fig 3.2. Metal distributions in Lake Burragorang sediments determined by sequential extractions

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Fig 3.3. Metal distributions in Lake Burragorang sediments determined by sequential extractions

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0 Percentage (%)Fraction of Mo DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7 Sites

120

100

80

60

40

20

0 Percentage (%) Fraction of Ni DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7 Sites Exchangeble Carbonate Fe-Mn Oxide Organic Inert

Fig 3.4. Metal distributions in Lake Burragorang sediments determined by sequential extractions

.

57 F

ig 3.5. ig Percentage (%) Fraction of Zn Percentage (%) Fraction of V Percentage (%) Fraction of Pb 100 120 100 120 100 120 20 40 60 80 20 40 60 80 20 40 60 80 0 0 0

sequential extractions sequential Metal distributions sediments LakeBurragorang in Metal W3DA W9DA2DA8DA9DA7DA5DA9M M7 M3 DWA39 DWA35 DWA27 DWA19 DWA18 DWA12 DWA9 DWA2 DWA3 W3DA W9DA2DA8DA9DA7DA5DA9M M7 M3 DWA39 DWA35 DWA27 DWA19 DWA18 DWA12 DWA9 DWA2 DWA3 W3DA W9DA2DA8DA9DA7DA5DA9M M7 M3 DWA39 DWA35 DWA27 DWA19 DWA18 DWA12 DWA9 DWA2 DWA3 Exchangeble

Carbonate 58 Sites Sites Sites Fe-Mn Oxide Fe-Mn determined by determined Organic Inert geochemistry. Metals of natural origin occur primarily in the residual sediment fraction [Jones and Turki, 1997; L’her Roux et al., 1998; Ianni et al., 2000]. Fe is largely present in the last fraction (the residual fraction), with moderate amounts associated with the Fe-Mn oxides fraction. The findings are similar as found by Korfalia and Davies [2004] and Takarina et al. [2004]. The site M7 at Werri Berri creek showed higher reducible phase than residual. The small concentration of Fe was also found attached to other remaining fractions.

The bar charts for Mn shows that it is mainly associated with exchangeable and carbonate phase from which it can be readily mobilised. Previous studies [Chakrapani and Subramanian, 1993; Ouddane et al., 1997] also concluded that the first two fractions are dominant phase for Mn. Speciation profile of Mo depicted the control of organic phase (79-90 %) over other on the location close to dam wall (DWA2, DWA3 and DWA9). The points other than that show Mo associated with inert phase (70-90%).

Except fraction one (1) Ni was extracted in all steps but largely hosted by residual fraction, which accounts for 28-58% and moderate affiliation with Fe-Mn oxides, organic and carbonate phases. These results are in agreement with the observations of Tessier et al. [1980] who suggested that a majority of the Ni in sediments was detrital in nature. Adamo et al. [1996] demonstrated that Ni in contaminated soils often occurs as inclusions within the silicate spheres rather than as separate grain. The Ni inclusions are protected against natural decomposition as well as reagent alteration, and only the dissolution of the silicates would ensure their extraction.

Speciation pattern of Pb suggests its strong association with residual fraction and moderate with Fe-Mn oxides and organic. There is also little Pb present in the carbonate form whereas exchangeable have very low or negligible presence of Pb. The results are in agreement with previous research [Jones and Turki, 1997; Akcay et al., 2003]. The distribution of vanadium is similar to Pb, being mostly dominated by inert fraction followed by reducible with minor amount in oxidisable. The high percentages of Zn total content in residual at all sites except DWA19, M3 and M7 where reducible and exchangeable fractions take over. Zn also appeared in the same pattern as Ni but Zn was also associated with exchangeable portion. The residual fraction accounting for 42.8 -48.2% of the total Zn concentration. This result is in 59 agreement with Baruah et al. [1996]; Stone and Droppo [1996]; Ma and Rao [1997]; Li et al. [2000]; Fytianos and Lourantou [2004]. Among the nonresidual fractions, the Fe–Mn oxide fraction was much more important than other fractions in all sediments, which accounted for 18–42% of total Zn. The Zn percentage bonded to organic and carbonates fraction are very similar. Small fraction of Zn (3-17%) also presents in exchangeable.

This is the first study that report metal speciation data for lake Burragorang sediments. With a few exceptions here and there, the speciation profile of a particular metal is same throughout the stretch of Lake Burragorang. The speciation patterns of As, Fe, Mo, Ni, Pb and V indicate their significant association with the residual fractions of sediments. Small percentage of Mo is hosted by first two phases mainly at upstream. Copper and Cr speciation demonstrated their high percentage association with residual and organic fraction, which make them least mobile. Substantial amount of metals like Cd, Co, Mn, and Zn are present in the first three fractions exchangeable, carbonate and reducible. The exchangeable and carbonate, which are considered to be weakly bound fractions and may equilibrate with the aqueous phase, thus become more bioavailable. The Fe-Mn oxide and the organic matter have a scavenging affect and may provide a sink for heavy metals. The release of the metals from this matrix will most likely be affected by the redox potential and pH. Moderate association of Ni and Pb in carbonate fractions and Fe-Mn oxide fractions thus has a possibility of becoming readily bioavailable. The total Fe in the sediments is quite high and even its lower amount bound to the exchangeable and carbonate fractions could cause deleterious effects.

Overall, data on the fractional distribution of heavy metals indicate that Cd, Co, Mn, and Zn have the highest migration mobility whereas Cu and Cr have least in Lake Burragorang sediments. The results showed the ease with which metals leach from sediments decreases in the order: Mn=Cd>Co= Zn > Ni > Mo> Pb>Fe>V> As>Cu>Cr.

60

Chapter IV. Distribution of heavy metals and their bioavailability using SEM and AVS in the sediments of Lake Burragorang

4.1 Introduction

The distribution of metals and their speciation in surfacial sediments of Lake Burragorang described in previous chapter generated interest to investigate the magnitude and variability of metals and nutrient levels in sediment cores, which will reflect the history of pollution events that have occurred over a span of decades.

Acid volatile sulphide and simultaneously extracted metal method is able to predict quite well the availability of various heavy metals for different organisms [Hoop et al., 1997]. The present chapter includes the AVS-SEM experiments conducted on selected cores to have better understanding of bioavailability of heavy metals.

River and freshwater sediments generally have higher organic contents than marine sediments. This leads to a rapid consumption of oxygen by aerobic decomposition of organic matter in sediments. As a result oxygen is depleted below a few millimeters of the sediment-water interface in freshwater sediments [Jorgensen and Sorenson, 1985].

In oxic sediments, the most important phases for metals are those containing hydrous iron and manganese oxides [Yu et al., 2001]. In anoxic sediments, sulphide phases dominate and the bioavailability of heavy metals is largely controlled by the absorption and coprecipitation of the metals with sulphide minerals. [Di Toro et al., 1990; DiToro et al., 1992; Ankley et al., 1996; Cooper and Morse, 1998].

Di Toro et al. [1992] gave a comprehensive description of the role of sulphide in the chemical activity of the metal in the sediment-interstitial water system. The system is characterised by treatment of the sediment with cold hydrochloric acid. The sulphide fraction released in this way is referred to as the acid volatile sulphide or AVS. AVS, comprising essentially iron monosulphides in sediments, are available for binding divalent cationic metals through the formation of insoluble metal-

61 sulphide complexes [Di Toro et al., 1990; Allen et al., 1993.; Huerta-Diaz et al., 1998], thereby controlling the metal bioavailability and subsequent toxicity for benthic biocommunities. The AVS-bound metals, with environmental concern (usually Cd, Cu, Ni, Pb and Zn), are extracted at the same time and are called simultaneously extracted metals (SEM). The ratio or the difference between AVS and SEM gives an indication of the potential sediment toxicity.

At ΣSEM/AVS < 1, sediments are interpreted as non-toxic because the reactive sulphide present exceeds the extractable sediment metal concentrations. Nevertheless, “non-toxic” sediments can also act as potential sources for heavy metal release to aquatic biota. This can turn into a consequential pollution source. Changing the aquatic conditions and exposing the anoxic sediment to an oxic environment can cause the sulphide material to be reoxidised and metals released to the water column [Delaune and Smith, 1985; Calmano et al., 1994; Petersen et al., 1997] The aquatic conditions can be changed through physical and chemical properties. These changes can occur by natural events such as storms, by human activity such as dredging or prop wash [Shipley et al.].

Since the mid 1990s, the AVS concept has been introduced in a number of risk assessment studies of anaerobic, heavy metal–polluted freshwater sediments. The AVS concept relies on the geochemical process of heavy metals being precipitated by an excess of sulphide as the result of their very low solubility products. Under anaerobic conditions the mobility of the metals is thus strongly reduced and toxic effects due to the presence of the metals are negligible [Buykx et al., 2002].

The significance of sulphide partitioning in controlling metal bioavailability in marine sediments spiked with cadmium was demonstrated by Di Toro et al. [1990, 1992]. Hoop et al. [1997] investigated the spatial and seasonal variations of acid volatile sulphide and simultaneously extracted metals in Dutch marine and freshwater sediments. AVS has been detected in 95% of the investigated sediment samples. The corresponding SEM/AVS ratio was found to be smaller than one in 19 out of 21 samples. According to literature data, toxic effects from heavy metals are expected to be absent under these conditions. This study has been used to examine the applicability of the AVS-concept in Dutch sediment quality criteria.

62 Grabowski et al. [2001] obtained SEM and AVS concentrations during spring and summer at six locations along Mississippi River floodplain. They found no spatial but temporal variation in SEM –AVS values. AVS concentrations were significantly greater during summer and spring. The sediments of Pearl River Estuary, China have been studied for AVS-SEM by Fang et al. [2005]. The results showed that AVS was not the most important phase for heavy metals in the surface sediments. However, AVS could play an important role in binding heavy metals in deep layer sediments of the estuary. They also compared sequential extraction procedure (SEP) and AVS-SEM measurements and suggested SEP can be used as an additional tool with the AVS method for assessing the potential bioavailability and toxicity of metals in sediment. Mackey and Mackay [1996] study on Mangrove Sediments, Brisbane River, Australia has shown a marked spatial variability in AVS and metal concentrations, and consequently bioavailability of metals. They mentioned the seasonal variations would further increase the observed variability in bioavailability. This variation should be taken into account when monitoring and assessing long- term trends in sediment.

In light of the importance of the AVS- SEM concept for assessing the ecological risk of metals in sediments it was considered necessary to evaluate the bioavailability of heavy metals in lake Burragorang and determine if there is any relationship between human land use pattern and AVS-SEM values

4.2 Study Area

Fourteen sediment cores were collected from different sites in Lake Burragorang during November 2002 (Fig 2.1) (details of sites are given in Table 4.1). The samples were processed as discussed in Chapter II and analysed for organic matter, carbonate, nutrients and heavy metals. The experimental methods have been described previously in Chapter II.

Four cores were collected in June 2003 for AVS-SEM study. (a) Site DWA2 (300 m upstream of Dam wall): This is an important location as most of the water is extracted near this point and supplied to Sydney Water, (b) DWA18: It is located at the inflow site of the Cox and Kedumba River (36 km upstream), (C) DWA27: It is located in the middle of lake and (d) DWA35: It is at the inflow site of the Nattai.

63 Table 4.1. Lake Burragorang monitoring sites

Station name Longitude Latitude Depth (AMG) (AMG) (m) DWA2 277157 6247707 88 DWA3 274394 6245774 87 DWA6 270817 6243290 80 DWA9 267645 6239826 75 DWA12 258408 6244960 47 DWA15 263883 6240307 65 DWA18 254049 6245232 33 DWA19 254187 6250243 10 DWA27 261877 6235614 54 DWA30 261718 6226644 36 DWA35 259190 6223714 29 DWA39 255717 6219495 12 MC3 275107 6244794 50 MC7 274338 6243280 11 UWS 13 254102 6218371 3 UWS 14 264197 6223203 10.3 UWS 15 261537 6230428 43.9

4.3 Results and Discussion

4.3.1 Organic Matter and carbonate content

The percentage of organic matter and carbonate content is given in Table 4.2 and depth profiles of changes are shown in Figs B1-B4. In general carbonate contents were more or less constant with a slight decrease at the bottom at all sites except DWA35.Organic matter decreases with depth on those sites, which are near to dam wall (DWA2, DWA6 and MC3). At sites near the inflow from Nattai river, i.e. DWA30 and UWS 14, a positive peak was observed at 10 cm. Interestingly down towards the Wollondilly River in south (DWA39 and UWS13) and Cox river in north (DWA12 and DWA18) organic percentage became relatively constant while going down. At DWA35 similar trend was observed for carbonate and organic percentage, however, two peaks were found at 10 and 25cm layers.

64 4.3.2 Nutrients

Nitrogen and phosphorus stored in the bottom sediments are a potential source of nutrients to the lake by internal loading. The extent to which this potential is realised depends to a larger extent on the oxygen content and pH of the water in and above the sediments. The SCA annual water quality monitoring report 2000-2001 stated elevated levels of nutrients and suspended solids during periods of wet weather [SCA, 2001a]. Higher concentration of particulate nutrients, suspended solids and metals were observed at lower layer (bottom) during floods [Sia, 2003]. Agricultural activities around the catchment could lead to increased levels of nitrogen and phosphorus. The present section discusses the scenario of nutrients in Lake Burragorang sediments.

Depth distribution of nutrients in the cores is displayed in Figs. B1-B4 and Table 4.2. Total phosphorus (TP) ranged from 60 (at UWS13) to 1360 (at DWA2) mg/kg and total nitrogen (TN) ranged from 314 (at DWA18) to 3769 (UWS14) mg/kg. The concentrations were generally higher at the top and decreased with depth. This observation is in agreement as reported by Provin et al. [1989] and Wang et al. [2004].

The nutrient profile observed at DWA2, DWA30, UWS13, UWS14, UWS15, DWA35 DWA39 and MC3 showed decrease in concentration with depth. Interestingly, at DWA6 a sharp decrease of nutrients was observed at 15 cm depth and TN and TP concentrations became very close to each other. After that erratic variation was observed in nutrient contents.

At DWA9, TP is almost constant in the top section until 20 cm thereafter a slight variation is observed, however, TN values decreased with a peak at 25cm. An irregular decrease in TP and TN was found at DWA15. Nutrients content at DWA27 increased until 20 cm (550 mg/kg; 1434 mg/kg) and then decreased beyond this depth. Somewhat constant TP behaviour observed at DWA12 whereas irregular increase found in TN concentrations.

The value of TP at DWA18 displayed variation throughout the depth range reaching ~1160 mg/kg at the bottom, which is higher compared to top slice value. On the

65 other hand, TN values decrease until 20 cm depth and then rise quickly beyond 20cm. The values at the bottom are bit higher than surface values.

Sediment-quality guidelines for nitrogen or phosphorus have not been established [Juracek, 2004]. TP concentrations at Lake Burragorang were found higher than Bellinger Estuary (TP 176 mg/kg) in northern New South Wales, which is considered to be almost pristine [Birch et al., 1999]. Nutrient concentrations in the bottom sediment varied substantially among the different sites. Most of them show positive trend (that is, nutrient concentration increased toward the top of the sediment core). These trends in nutrient concentrations may be related to an increase in fertilizer use, livestock production and sewage-treatment plants around the catchment. Alternatively, the trends may be indicative of diagenesis (that is, post depositional changes in the sediment caused by various processes including decomposition) [Juracek, 2004].

4.3.3 Background and Metal Data

The background levels of metals in sediments of Lake Burragorang have not yet been reported. It is necessary to establish natural background metal levels before the extent, if any, of heavy metal contamination can be estimated. Such background levels are subtracted from the total values to yield an estimate of the anthropogenic contribution. Background levels can be estimated by: (a) Average metal concentrations of texturally- equivalent sediments reported in the literature. (b) Direct measurements of metal concentrations in recent texturally and mineralogically- equivalent sediments from a known pristine region. (c) Direct measurements of metal concentrations in texturally-equivalent sub-surface core samples obtained from a depth below any possible contamination or biological mixing [Loring and Rantala, 1992]. Third approach is commonly used to determine the preanthropogenic element values [Peterson et al., 1990; Valette-Silver, 1993; Murray, 1996; Birch and Taylor, 1999]. Generally metal levels are irregular and high near the top of the core, values decline down the core to a constant levels upto the base of the core.

66

Table 4.2. Spatial and vertical distributions of carbonate content, organic matter and nutrients in sediment cores of Lake Burragorang

S.No. Sample ID Depth TP TN Organic Carbonate Moisture (cm) (mg/kg) (mg/kg) matter (%) (%) (%) 1 DWA2 5 1360 2500 12.3 3.3 75.9 2 10 831 1818 11.5 3.0 73.6 3 15 1350 2007 10.1 2.9 72.6 4 20 1203 1522 9.4 2.6 68.6 5 25 1111 1424 9.2 2.1 66.4 6 30 919 1557 7.8 2.0 62.1 7 35 875 610 21.4 5.0 43.2 8 40 616 815 7.2 2.0 66.2 9 DWA6 5 1110 1626 10.7 2.6 74 10 10 1131 1318 10.3 2.5 73.5 11 15 317 339 9.2 2.3 71.3 12 20 948 1485 9.2 2.5 74.6 13 25 929 1161 9.0 2.3 70.4 14 30 760 1373 8.5 2.1 67.4 15 35 645 1590 8.6 2.2 67.3 16 40 602 1505 2.5 0.6 58.2 17 DWA9 5 680 2150 9.7 2.6 65.2 18 10 595 1518 8.6 2.6 71.2 19 15 563 1542 7.8 2.2 64.7 20 20 527 1539 8.3 2.3 69.6 21 25 393 1934 10.1 2.4 53.9 22 30 491 1758 13.6 1.5 55.9 23 DWA15 5 587 1895 8.5 3.1 68.5 24 10 564 1469 8.1 3.1 68 25 15 879 2160 5.3 5.3 67 26 20 369 843 5.2 4.9 66.4 27 25 492 1348 5.7 4.3 59.4 28 30 358 820 5.5 4.7 56.2 29 35 461 1530 11.3 1.7 54.4 30 DWA27 5 233 429 11.8 2.9 71.3 31 10 368 627 11.3 2.7 70.8 32 15 532 1435 10.5 2.2 61.2 33 20 550 1558 10.9 2.2 69.6 34 25 368 1166 8.9 1.3 61.9 35 UWS15 5 729 2135 7.3 1.4 69.6 36 10 674 1921 14.2 2.9 71.3 37 15 549 1561 10.3 1.9 71.6 38 20 525 1437 14.9 2.5 69.3 39 25 516 1506 11.4 2.1 61.2 40 30 500 1436 15.8 2.7 70.4 41 DWA30 5 632 2306 6.9 1.4 69.4 42 10 575 2225 13.5 2.7 73 43 15 502 1725 7.6 1.5 70.2 44 20 340 1145 8.1 1.4 66.2 45 25 300 1278 7.5 1.1 57.8 46 30 297 1176 5.3 0.9 59.8

Continued--- 67

S.No. Sample ID Depth TP TN Organic Carbonate Moisture (cm) (mg/kg) (mg/kg) matter (%) (%) (%) 47 UWS14 5 867 3746 4.3 0.5 80.9 48 10 1173 3769 34.4 3.5 83.9 49 15 909 3099 21.8 2.9 80.8 50 20 806 1757 11.5 1.9 73.6 51 25 972 1746 12.3 1.8 71.5 52 DWA35 5 648 2321 3.5 1.2 65.8 53 10 623 1710 14.7 14.5 67.6 54 15 592 1684 5.9 4.0 73.5 55 20 486 1230 6.0 4.2 65.6 56 25 647 1181 11.8 7.8 58.6 57 30 364 1067 9.8 2.3 63.6 58 35 302 958 5.2 1.0 64.9 59 DWA39 5 361 1681 14.4 2.3 65.3 60 10 415 1622 7.8 1.6 64 61 15 425 1318 7.6 1.6 63.5 62 UWS13 5 336 1593 7.6 1.1 69.5 63 10 350 1602 8.7 1.5 71 64 15 347 1394 7.5 1.4 69.5 65 20 380 1248 7.1 1.2 52.6 66 25 60 353 6.8 1.3 62.8 67 30 214 600 8.2 1.2 65.5 68 DWA12 5 572 2421 9.6 1.7 72.5 69 10 613 1475 8.0 1.5 63.6 70 15 604 1333 7.5 2.3 69.8 71 20 591 1754 8.0 1.3 62.3 72 25 720 3100 14.5 1.2 63.4 73 30 743 3482 16.3 1.4 67.1 74 35 733 3169 16.2 1.4 64.9 75 DWA18 5 1132 1613 10.2 1.9 75.4 76 10 699 2484 9.1 1.5 66.7 77 15 633 2104 9.4 1.5 66.6 78 20 150 315 9.8 1.6 72.8 79 25 585 1623 9.3 1.7 66.1 80 30 591 1906 10.2 1.5 62.9 81 35 1160 2177 10.5 2.2 70.3 82 MC3 5 682 2508 15.7 3.0 74.6 83 10 686 2397 17.1 2.9 74.6 84 15 593 2068 10.2 2.5 76.1 85 20 591 1626 8.5 5.2 72.7

68 Advantages of the background approach are that it requires a minimum of field data and no quantitative toxicity assessments. The disadvantages are that it may be difficult both to find suitable reference sites and to determine what levels are acceptable "background", presumably non-toxic concentrations. This approach may also result in pollutant levels that are lower -- perhaps even far lower -- than are toxic to benthic organisms [Batts and Cubbage, 1995]. Eventhough, the method is widely used to discern the natural presence and the anthropogenic contribution since it is the simplest and most straightforward of the guideline development methods.

In order to assess the background levels in Lake Burragorang sediments only those sites and metals were considered which shows constant levels down the core (Table 4.3 and Figs. C1-C5). The metals with irregular trend were not selected to estimate background levels (Table 4.4). The regular trend was observed at around 30-45 cm of core depth. Based on analysis background concentration were established as 4.7, 0.2, 23, 12, 20, 29000, 22, 660, <0.2, 0.25, 19.7, 0.13, 37 and 68 mg/kg for As, Cd, Cr, Co, Cu, Fe, Pb, Mn, Hg, Mo, Ni, Se, V and Zn, respectively. The background levels are quite comparable to other studies (Table 4.5).

4.3.4 Metals

The concentrations of metals at different location and depths are displayed in Table 4.3 and graphical presentation is given in Figs C1-C5. The distribution of trace metals is highly variable. The dominant metal are Fe and Mn followed by Zn, V and then Cr, Pb, Ni, Cu, Co and As. The other metals (Cd, Mo and Se) present in lesser amounts and, at few sites, were closer to the detection limit (Table 4.3). Hg was below detection limit in all locations. Iron and Mn concentrations have been plotted separately to get better understanding of their trends.

At DWA2 the concentrations of all metals (Fig C1 and Table 4.3) were relatively uniform with positive and negative peaks at 20 and 35 cm, respectively, whereas Fe and Mn profile showed an increase at 10 cm and decrease at 35 cm. A sharp decrease of Zn concentrations was observed at 10 cm depth at DWA6, however, beyond that no significant difference was observed with depth. Concentrations of other metals moderately decreased with depth, but with a peak at 10 cm. No regular trend was observed in Fe profile. At DWA9 metal concentrations except Mn increased to a

69 maximum value at 15 cm depth and then decreased to a relatively constant value. Mn profile showed continuous decrease in value with increasing depth.

A regular decrease was observed with depth at DWA15 for all metals except Zn and Fe. Zn and Fe also decreased but with positive peak at 15 cm. At DWA27 rapid increase was noticed in metal profile between 15 to 20 cm but overall concentrations decreased with increasing depth. At UWS15, located just in the middle of lake, it was found that the metal concentration decreases with depth. Metals profile at DWA30 was found similar to DWA 9 profile except a slight escalation in Pb, V and Zn concentration at 25 cm depth.

More or less constant profile was observed throughout the depth for metals except Fe and Mn at UWS14. Metal concentrations at DWA35 follow the trend of DWA9 and DWA30. Only 15 cm long core was collected at site DWA39 and concentrations of Cd, Co and Ni were observed constant whereas showed variation around 10cm. All metals concentration including Fe and Mn at UWS13 decreased with increasing depth until 20 cm and then shows an increase. At DWA12 Cd, Co, Cu and Ni pattern are more or less same through out the core. Cr, Mo, Pb and V profile moderately decrease down the core length. Zn behaviour also decreases but with a peak at 20 cm. Substantial decrease was found in Fe and Mn concentrations from 5 to 10 cm, however, afterwards moderate decrease was observed. Concentrations/depth profile at DWA18 in general showed decrease in concentrations with slight variation at 20 cm. whereas Fe and Mn trend appeared irregular. Near Weeri Berri creek (MC3) concentrations of all metals were almost constant down the depth. Fe and Mn pattern were noticed with a sharp increase and then a decrease with depth.

Almost all metal concentrations near damwall and around middle of lake were substantially elevated over background levels estimated for lake Burragorang sediments. Manganese at DWA30, UWS14, DWA35, UWS13 and DWA18 was observed below the background limits (660mg/kg). Arsenic, Cu, Pb, Zn and Fe were below the background in all segments of core UWS 13. Sediments of Lake Burragorang results show clear indication of heavy metal accumulation and the contamination is potentially significant upstream of the lake.

70

Table 4.3. Variation in metal concentrations with depth in sediment core samples

S.No. Sites Depth Metals (mg/kg) (cm) As Cd Cr Co Cu Pb Hg Mo Ni Se V Zn Mn Fe 1DWA25 5.60.2733132529ndnd 25 0.76 58 92 1090 38100 2 10 6.6 0.30 31 14 27 29 nd 0.13 25 0.69 56 96 1280 48700 3 15 6.4 0.28 34 13 23 29 nd nd 27 0.74 53 92 980 39200 4 20 8.2 0.33 37 17 31 33 nd 0.15 32 0.68 58 120 900 35600 5 25 5.8 0.20 31 15 24 24 nd nd 27 0.46 46 89 880 34500 6 30 6.7 0.20 31 17 26 26 nd nd 29 0.45 43 97 820 35000 7 35 4.9 0.15 23 12 21 17 nd nd 23 0.52 30 74 710 29700 8 40 5.5 0.16 36 17 26 27 nd nd 27 0.16 51 92 980 35700 9DWA65 7.30.2630142926ndnd 30 0.42 52 180 1820 39200 10 10 11 0.33 38 17 32 35 nd nd 32 0.29 66 110 1380 41300 11 15 10 0.31 39 17 30 34 nd 0.27 33 nd 63 110 1290 43200 12 207.70.3333142429nd0.1427nd5393121042100 13 258.30.2933152628nd0.2327nd5395110042400 14 307.00.2232152627ndnd 27 nd 49 94 1020 36100 15 356.60.1831162426ndnd27nd4786116036300 16 407.10.2131162926ndnd 28 0.12 42 94 940 36500 17 DWA9 5 8.4 0.30 31 17 26 28 nd nd 26 nd 53 96 2860 40400 18 10 8.4 0.33 32 15 26 28 nd 0.44 28 nd 52 100 1350 41100 19 15 10 0.30 40 20 31 33 nd 0.65 35 nd 66 120 1140 42600 20 206.50.2336172629ndnd28nd5292101034900 21 25 6.3 0.22 27 12 21 25 nd 0.49 21 nd 40 69 730 26100 22 30 5.3 0.22 24 13 20 24 nd 0.59 20 nd 36 68 600 25500 23 DWA15 5 8.3 0.28 32 16 26 28 nd 4.0 26 nd 55 91 2520 40300 24 107.80.2431142426nd0.4425nd5182118035600 25 157.80.2631142426nd0.2926nd5090100038300 26 206.60.1830142325ndnd26nd458387034600 27 256.00.1630142323nd0.1424nd4478106033800 28 304.50.2227142221ndnd22nd397381028400 29 354.60.2324142221ndnd 21 nd 32 69 1060 25000 30 DWA27 5 11 0.23 32 16 25 28 nd nd 26 nd 54 85 2460 45300 31 108.10.2031142227ndnd25nd5179139038400 32 156.80.1729142026ndnd24nd466997033900 33 20 11 0.27 53 24 33 44 nd nd 39 nd 88 130 2150 54000 34 255.40.1729121726ndnd 19 nd 47 61 800 32000 35UWS155 8.70.2137182635ndnd31nd6990180047200 36 109.80.2533162330ndnd25nd5679145046700 37 15 8.2 0.22 34 14 22 28 nd 0.65 25 nd 56 78 920 42800 38 207.50.2432152127ndnd24nd5372138038000 39 255.40.1630131927ndnd22nd476870031900 40 304.30.2434122028ndnd 23 nd 50 67 600 31500 41DWA305 6.40.1729142026ndnd22nd487255036900 42 105.80.1729121926ndnd21nd476648034700 43 156.80.4233152227ndnd24nd568250037100 44 205.30.1729121725ndnd19nd507040029200 45 256.10.1825101630ndnd17nd547544028000 46 305.50.3024111822ndnd 19 nd 49 69 440 28800

Continued---

71

S.No. Sites Depth Metals (mg/kg) (cm) As Cd Cr Co Cu Pb Hg Mo Ni Se V Zn Mn Fe 47 UWS14 5 4.7 0.18 25 12 24 25 nd 0.16 22 nd 42 78 290 35700 48 104.50.1825112523ndnd22nd417825032400 49 15 5.1 0.18 26 13 27 25 nd 0.16 23 nd 43 77 330 38200 50 20 4.6 0.15 25 13 24 23 nd 0.23 23 nd 43 67 220 34600 51 25 5.1 0.15 25 13 25 23 nd 0.21 23 nd 47 74 220 33300 52DWA355 6.70.1732152027ndnd22nd527571039800 53 106.50.2129131826ndnd20nd476654036700 54 157.90.1933152331ndnd23nd557758041400 55 206.50.2032142028ndnd23nd527247037200 56 254.90.1226121621ndnd18nd456044029200 57 304.50.1029121619ndnd 18 nd 47 59 450 28700 58 354.50.1727111523ndnd 17 nd 46 54 470 29900 59DWA395 2.80.1224111417ndnd17nd415735026300 60 104.60.1229121720ndnd 19 nd 49 65 380 29600 61 155.50.1329121722ndnd 19 nd 52 63 340 29700 62UWS135 2.00.1028131719ndnd19nd476433025200 63 102.60.5229141820ndnd21nd516525025500 64 153.60.1028131719ndnd20nd516325025800 65 20 2.2 ND 20 9.3 12 14 nd nd 15 nd 36 45 200 19100 66 253.2ND34151818ndnd 26 nd 51 62 320 26700 67 30 4.2 0.34 29 14 19 19 nd 0.18 24 nd 51 72 360 26000 68DWA125 9.20.2326152525nd0.2923nd4786329041700 69 10 7.0 0.23 25 12 22 22 nd 0.3 21 nd 39 82 1420 34000 70 156.80.2024132423nd0.2123nd3887106035200 71 20 6.3 0.20 23 15 23 21 nd 0.18 24 nd 34 110 1040 31100 72 255.90.2019142420ndnd22nd2265102030800 73 305.90.2320162720ndnd 24 0.12 23 72 930 26600 74 355.80.2420152918ndnd 24 0.24 22 72 720 24100 75DWA185 6.10.2624122521ndnd19nd397942035100 76 104.90.2921112521ndnd18nd408651038500 77 15 3.8 0.59 18 9.6 28 21 nd nd 13 nd 37 81 380 26000 78 204.90.4023102623ndnd17nd398738029600 79 254.70.3524132423ndnd21nd358652028200 80 303.70.3021112519ndnd 20 nd 27 84 410 25100 81 354.00.2921122319ndnd 22 nd 27 83 660 28100 82 MC3 5 7.5 0.19 22 15 20 26 nd nd 17 nd 42 73 970 44700 83 106.80.2123152128ndnd17nd4579421068000 84 157.30.1924152128ndnd 18 nd 45 74 790 41500 85 206.50.1825142025ndnd 18 nd 45 71 740 47400 MDL 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.5 0.1 0.5 0.5

nd = Not detected MDL = Method detection limit

72

Table 4.4. Background metal levels for Lake Burragorang from sedimentary metal concentrations

Station Metal (mg/kg) As Cd Co Cr Cu Fe* Hg Mn Mo Ni Pb Se V Zn DWA2 4.9-8.2 0.15-0.33 12-17 23-37 21-31 2.9-48 <0.1 710-1280 0.13-0.15 23-32 17-33 0.16-0.8 30-58 74-120 DWA6 7-8.3 0.18-0.33 14-17 31-39 24-32 3.6-4.3 <0.1 940-1820 0.14-0.27 27-33 26-35 0.12-0.4 42-66 86-180 DWA9 5.3-10 0.22-0.33 12-20 24-40 20-31 2.6-4.2 <0.1 600-2860 0.44-0.65 20-35 24-33 <0.1 36-66 68-120 DWA15 4.5-8.3 0.16-0.28 14-16 24-32 22-26 2.5.-4.0 <0.1 810-2520 0.14-0.44 21-26 21-28 <0.1 32-55 69-91 DWA27 5.4-11 0.17-0.27 12-24 29-53 17-33 3.2-5.4 <0.1 800-2460 <0.1 19-39 26-44 <0.1 46-88 61-130 UWS15 4.3-9.8 0.16-0.25 12-18 30-37 19-26 3.1-4.7 <0.1 600-1800 0.65 22-31 27-35 <0.1 47-69 67-90 DWA 30 5.3-6.8 0.17-0.42 10-15 24-33 16-22 2.8-3.7 <0.1 400-550 <0.1 17-24 22-30 <0.1 47-56 66-82 UWS14 4.5-6.1 0.15-0.18 11-13 25-26 24-27 3.2-3.8 <0.1 220-330 0.16-0.23 22-33 23-25 <0.1 41-47 67-78 DWA35 4.5-7.9 0.1-0.21 11-15 26-33 15-23 2.9-4.1 <0.1 440-710 <0.1 17-23 19-31 <0.1 45-55 54-77 DWA39 2.8-5.5 0.12-0.13 11-12 24-29 14-17 2.6-3.0 <0.1 340-380 <0.1 17-19 17-22 <0.1 41-52 57-65 UWS 13 2-4.2 0.1-0.52 9.3-15 20-34 14-20 1.9-2.7 <0.1 200-360 0.18 15-26 12-19 <0.1 36-51 45-72 DWA12 5.8-9.2 0.2-0.24 12-16 19-26 22-29 2.4-4.1 <0.1 720-3290 0.15-0.3 21-24 18-25 0.12-0.2 22-47 65-110 DWA18 3.7-6.1 0.26-0.59 9.6-13 18-24 23-28 2.5-3.9 <0.1 380-660 <0.1 13-22 19-23 <0.1 27-40 79-87 M3 6.5-7.5 0.18-0.21 14-15 22-25 20-21 4.1-6.8 <0.1 740-4210 <0.1 17-18 25-28 <0.1 42-45 71-79 Background Level 4.7 0.2 12 23 20 2.9 <0.1 660 0.25 19.7 22 0.13 37 68

*Fe in % weight. Metals shown in italic not selected for assessing background levels

Table 4.5. Background metal levels for Lake Burragorang with other matrices

Rural Georges Shallow reach Metals Lake Hawkesbury River/Port Coxs Sydney Crust water South (mg/kg) Burragorang Rivera Hackingb Riverc Habourb abundanced Sedimentd sedimente Creekf Shaleg Others h As 4.7 ------3 Cd 0.2 - 1 1 2 0.11 0.17 - - 0.3 Co 12 - 5 20 16 20 14 13 - 19 Cr 23 - - 30 - 100 72 60 - 90 Cu 20 18 9 45 10 50 33 56 31 45 Fe* 2.9 2.2 2.5 4 3.9 4.1 4.1 6.5 - 4.7 i Hg <0.2 ------0.01-0.24 Mn 660 - 56 1000 131 950 770 850 - 850 Mo 0.25 ------Ni 19.7 - - 50 26 80 52 35 23 68 Pb 22 22 32 25 33 14 19 22 28 20 Se 0.13 ------V 37 - - - 60 - - - - - Zn 68 57 43 130 47 75 95 92 65 95

*Fe in %. Weight. (-) No establish data a [Shotter et al., 1995] b [Irvine and Birch, 1998] c [Siaka, 1998] d [Bowen, 1979] e [Wedepohl, 1970] f [Thomas and Thiel, 1995] g [Turekian and Wedepohl, 1961] h [PTI, 1989] i [Syers et al., 1973]

73

4.3.5 Acid Volatile Sulphide and Simultaneously Extracted Metals

The AVS and SEM (Cd, Cu, Ni, Pb and Zn) concentrations and SEM/AVS ratio of sediment samples are shown in Table 4.6. The highest sulphide levels were obtained from site DWA2 (range from 0.59 to 0.12 μmol/g), while lowest levels were obtained from site DWA35 (range from 0.25 to 0.09 μmol/g). The distribution of AVS with depth in the sediment cores is presented in Fig 4.1. No regular trend was observed in the AVS pattern of the cores. In general, most cores had low AVS contents at the surface, higher at intermediate depth and low towards the bottom of the core. Two positive peaks of AVS were found at 10 cm and 20 cm layers and 10 cm and 25 cm in cores from DWA2 and DWA35, respectively. In core DWA18, two peaks at 15 cm and 25 cm were identified. Only one peak at 25 cm was found in core DWA27. From the above distribution patterns, it appears that the AVS contents generally decrease along the profile with peaks at various depths.

In all the sites among HCl-extractable metals (SEM) Cd concentrations were lowest and Zn was highest. Copper, Ni and Pb were intermediate. Vertical distribution of SEM is shown in Fig 4.1. No variation was found in Cd concentration with depth in all the four sites. Concentrations of Cu, Ni and Pb at all stations were more or less same at all depths. Zn concentration generally decreased with depth except few peaks in cores DWA2, DWA27 and DWA35.

The results showed that these simultaneously extracted metals at all stations were higher than AVS and their ratio was found greater than 1, which indicates that available AVS is not sufficient to bind with the extracted metals. This reveals that AVS is not a major metal binding component for Lake Burragorang sediments and contained metals are potentially bioavailable to benthic organisms.

The levels of AVS concentrations measured in the Lake Burragorang sediments is low compared to values reported in the literature for fresh water sediments [Machesky et al., 2004]. AVS concentration depends on season and depth. Many researchers observed variation in AVS levels with season [Aller, 1977; 2001; Grabowski et al., 2001; Machesky et al., 2004; Morse and Rickard, 2004]. Major decrease occurred in the winter because in colder temperatures, FeS formation rates 74

Table 4.6. Concentrations of AVS and SEM alongwith depth in sediments of Lake Burragorang

Depth AVS Cd Cu Pb Ni Zn ΣSEM SEM/AVS (cm) μmol/g μmol/g μmol/g μmol/g μmol/g μmol/g μmol/g

DWA2 5 0.50 0.0013 0.075 0.077 0.065 0.34 0.56 1.1 10 0.59 0.0017 0.086 0.087 0.075 0.35 0.60 1.0 15 0.28 0.0009 0.055 0.068 0.06 0.29 0.47 1.7 20 0.45 0.0018 0.071 0.063 0.053 0.28 0.46 1 25 0.29 0.0011 0.046 0.058 0.053 0.21 0.37 1.3 30 0.12 0.0011 0.071 0.053 0.039 0.14 0.30 2.6 35 0.18 nd 0.077 0.045 0.037 0.14 0.30 1.7 DWA18 5 0.35 0.0018 0.057 0.053 0.065 0.26 0.44 1.2 10 0.38 0.0017 0.035 0.053 0.06 0.31 0.45 1.2 15 0.50 0.0030 0.039 0.063 0.082 0.29 0.48 1 20 0.30 0.0019 0.049 0.042 0.044 0.23 0.37 1.2 25 0.40 0.0016 0.088 0.044 0.056 0.40 0.59 1.5 30 0.13 0.0013 0.069 0.048 0.049 0.31 0.47 3.7 DWA27 5 0.31 0.0014 0.097 0.087 0.12 0.38 0.69 2.2 10 0.17 0.0011 0.074 0.087 0.107 0.37 0.64 3.7 15 0.18 0.0011 0.066 0.072 0.082 0.28 0.50 2.8 20 0.20 0.0013 0.058 0.072 0.078 0.25 0.45 2.3 25 0.33 0.0012 0.088 0.077 0.077 0.28 0.52 1.6 30 0.26 nd 0.049 0.063 0.058 0.25 0.41 1.6 DWA35 5 0.21 0.0012 0.072 0.068 0.073 0.28 0.49 2.3 10 0.25 0.0012 0.052 0.063 0.058 0.25 0.42 1.7 15 0.14 0.0013 0.077 0.082 0.072 0.31 0.54 3.9 20 0.10 0.0015 0.047 0.047 0.043 0.18 0.32 3.1 25 0.17 nd 0.077 0.068 0.066 0.28 0.49 2.9 30 0.09 nd 0.036 0.042 0.034 0.18 0.30 3.4 35 0.10 0.0012 0.035 0.042 0.036 0.14 0.25 2.5

75

Table 4.7. Guidelines for determining metal toxicity to benthic organisms in freshwater sediments (values in mg/kg) [Grabowski, 2001]

Metal Threshold effects Upper effects level (TEL) threshold (UET)

Cd 0.596 3 Cu 35.7 86 Pb 35 127 Ni 18 43 Zn 123.1 520

were lower and a smaller supply of FeS-rich particles was brought up from below by bioturbation [Aller, 1977].

For the present study, sediments for SEM and AVS were collected for analysis in winter season and hence the low level of AVS was found. Low AVS concentrations indicate that most metals are bound by sediment constituents other than AVS [Allen et al., 1993]. A study conducted by Lawrence Berkeley National Laboratory (LBNL) also had sites with SEM-AVS values greater than one due to relatively low AVS values and not necessarily high concentrations of metals. No toxicity to benthic organisms was observed from these LBNL sites [Grabowski et al., 2001]. Other constituents in the sediment, such as iron and manganese oxides and organic matter, may have decreased the bioavailability of heavy metals [Di Toro et al., 1990]. The unbound metals toxicity to benthic organisms can be explained by analyzing individual SEM concentrations according to their upper effects threshold (UET) levels (Table 4.7). All the locations had individual SEM concentrations lower than their UET. Even though these investigated metals were bioavailable in the sediment their individual metal concentrations are not expected to be toxic to benthic organisms.

76 SEM Profile at DWA2 AVS Profile at DWA2

SEM (umol/g) AVS (umol/g) 0.00.10.20.30.4 0.00 0.20 0.40 0.60 0.80 0 0 5 5 10 Cd 10 15 Cu 15 (cm)

h (cm) h 20 Pb 20 AVS 25 Ni 25 Depth Dept 30 Zn 30 35 35 40 40

SEM Profile at DWA18 AVS Profile at DWA18

SEM (umol/g) AVS (umol/g) 0.0 0.1 0.2 0.3 0.4 0.5 0.00 0.10 0.20 0.30 0.40 0.50 0.60 0 0 5 5 10 Cd 10 Cu 15 15 Pb AVS 20 20 Ni Depth (cm) 25 Depth (cm) 25 Zn 30 30 35 35

SEM Profile at DWA27 AVS Profile at DWA27

SEM (umol/g) AVS (umol/g) 0.0 0.1 0.2 0.3 0.4 0.5 0.00 0.10 0.20 0.30 0.40 0 0 5 5 10 Cd 10 Cu 15 15 Pb AVS 20 20 Ni Depth (cm) 25 Depth (cm) 25 Zn 30 30 35 35

SEM Profile at DWA35 AVS Profile at DWA35

SEM (umol/g) AVS (umol/g) 0.0 0.1 0.2 0.3 0.4 0.00 0.10 0.20 0.30 0 0 5 5 10 Cd 10 15 Cu 15 20 Pb 20 AVS 25 Ni 25 Depth (cm) Depth (cm) Depth 30 Zn 30 35 35 40 40

Fig 4.1. AVS and SEM distribution with depth

77

Chapter V. Sedimentary record of heavy metal pollution of Lake Burragorang using 210Pb dating

5.1 Introduction

Knowledge of the formation and history of a lake is important from the point of understanding its structure and is also vital for its management. Sediments provide a history of our environmental misdeeds. Sediments deposited within aquatic environments are principally derived from weathering processes (eg. erosion, abrasion), with major transportation from terrestrial sources under high runoff from storms and floods. In addition, discharges from urban, industrial and mining activities are potential sources of particulates. Anthropogenic contaminants, including metals, organics and nutrient are associated with particulate and dissolved inputs to natural waters. [Arakel, 1995; ANZECC, 2000]. Particulate material entering the aquatic system is held in suspension until it is deposited and incorporated into the base sediments. A key issue that has affected aquatic environment is sedimentation. The rate of sedimentation and the change in rate of sedimentation are two of the most important parameters to interpret the depositional history and health of coastal environments. Sediment dating is used to calculate sedimentation rate and accumulation rate for different substances. The distribution of a variety of substances in annually deposited sediments has been used to provide information on pollution chronologies and paleoenvironments. If the concentration of substances at different depths is compared with the corresponding ages, the accumulation rates of these substances at different times may be determined from the sediment accumulation rate. This gives a better picture of the historical development in a given lake or marine area. Among different variables metals have received special attention because they are persistent so uncertainty introduced by compound degradation is eliminated. Furthermore, they can pose ecotoxicological risks at low concentrations [Benoit and Rozan, 2001].

78 5.2 Lead –210 Radiometric Dating

One of the most promising methods of dating on a time scale of 100-200 years is by means of 210Pb, a natural radioisotope with a half-life of 22.26 years. This is especially useful here in Australia as it can determine environmental impact since European settlement [Harrison, 2003]. The development of this technique was first initiated by Goldberg [1963], and it was first applied to the dating of lake sediments by Krishnaswamy et al. [1971]. 210Pb has been developed for a number of applications, including the depositional rates of sediments in lakes [Oldfield and Appleby, 1984] ), river floodplains and reservoirs [Owens et al., 1999], through to the dating of Antarctic snow [Lambert and Sanak, 1989] and cave deposited spelcothems [Bierman et al., 1998]. The technique has also been used to understand the impact of European settlement on terrestrial and aquatic ecosystem by analysing and dating pollen, charcoal, diatom, chironomid and inorganic content on Australian sediments [Colliton, 2001; Agnew, 2002; Haberle et al., 2006].

Lead-210 is a member of the uranium-238 decay series and is produced by the decay of the intermediate isotope 226Ra (half life 1622 yrs) to the inert gas 222Rn (half life 3.83 days) followed by a series of short lived isotopes to 210Pb [Brenner et al., 1994].

The 210Pb accumulates in lake and river sediments via a number of different pathways- erosion, wash-in and atmospheric dropout all contributes to effectively concentrate the amount of 210Pb. The 210Pb present in sediments is described and analysed as two components, 'supported' 210Pb and 'unsupported' 210Pb (Fig 5.1).

226Ra in the sediment within the catchments area enters via erosion wherein it decays to 210Pb. The 210Pb formed by the 'in situ' decay of 226Ra is called the 'supported 210Pb. The supported 210Pb is normally assumed to be in radioactive equilibrium with the radium, however, in the natural system this equilibrium is disturbed by a supply of 210Pb from other sources. Three components are identified by which excess 210Pb reaches the sediments. The first and second routes are due to atmospheric fallout. 222Rn is formed within the soil in the catchment area and being a gas, it escapes and diffuses into the atmosphere [Harrison, 2003]. The 222Rn in the atmosphere then decays to 210Pb, which attaches to airborn particulate matter and either – (A) falls directly into the catchment or (B) falls in the catchment region and is washed in by

79 rain or erosion. In the third route (C) 222Rn escapes up the water column due to the decay of 226Ra already in the river bed. Part of the 222Rn will migrate up to the surface of the water and escape to the atmosphere, where it will decay to 210Pb and some will decay to solid matter before reaching the water- atmosphere interface and return to the river bed. Lead-210 activity (components A, B and C) in excess of the supported activity is called the ‘excess’ or unsupported 210Pb. The total amount of 210Pb in a particular system is the total 210Pb supplied by both the 'supported' and 'unsupported’.

Fig 5.1. Pathways by which 210Pb reaches lake sediments [Oldfield, 1981; Organo, 2000]

Unfortunately, the activity of 210Pb cannot be measured directly as it is a beta emitter and peaks on the spectrum is difficult to distinguish due to substantial background. Instead, 210Ra and 210Po are analysed, as they are alpha emitters. Alpha emitters tend to show sharper peaks on a spectrum [Harrison, 2003]. By definition, the activity of 226Ra is in equilibrium with the 'supported 210Pb, and the activity of 210Po is assumed to be in equilibrium with the total 210Pb. The unsupported 210Pb activity is determined from it's granddaughter isotope Polonium-210, which is assumed to be in secular

80 equilibrium with 210Pb [Heijnis et al., 1987; Ivanovich et al., 1992; Ravichandran et al., 1995]. The activity of the 'unsupported 210Pb in a given sample is found by subtracting the activity of 226Ra from the activity of the 210Po. The 'unsupported 210Pb is generally used in calculations to determine sedimentation rate of a particular system [Oldfield and Appleby, 1984; Harrison, 2003].

5.3 Models for Sedimentation Rate Determination

The 210Pb dating technique can be applied to determine sedimentation rates and age profiles through the use of modeling. There are two main models used for age determinations using the 210Pb dating method [Organo, 2000]. The first of these is the constant rate of 210Pb supply or CRS model, which assumes that the supply of 210Pb to the accreting material is occurring at a constant rate. In this model the initial unsupported 210Pb activity varies inversely with the mass accumulation rate [Appleby and Oldfield, 1992]. The second model has been termed the constant initial concentration model or CIC model. This model assumes that the initial activity of unsupported excess lead-210 is the same at all depths in the core independent of the sedimentation rate [Geyh and Schleicher, 1990]. It is widely accepted that the atmospheric deposition of 210Pb in any region is governed by local geographical or meteorological factors, and is reasonably constant when averaged over several years. It is then reasonable to suppose that there will be a constant rate of accumulation of unsupported 210Pb, and that each layer of sediment will have the same initial unsupported 210Pb concentration [Appleby and Oldfield, 1992]. The CIC model has been applied within this study, however, the two models yield the very similar results if the accumulation rates are constant and not too large [Oldfield, 1981; Chanton et al., 1983; Appleby, 1993; Turner and Delorme, 1996] .

The current study has been undertaken to study the variability in metals and nutrients concentrations through lead-210 dating within lake Burragorang and compared with past data of rainfall and bushfire.

5.4 Sampling Locations

Fourteen sediment cores were collected from different locations (Fig 1.1) of lake Burragorang to study different variables (Chapter IV). Out of sixteen cores, three cores were selected (DWA2, DWA18, DWA35) to perform sedimentation rate study 81 at Australian Nuclear Science and Technology Organisation, Sydney (The sedimentation rate could not be performed on all fourteen sediment cores due to financial constraints).

5.5 Selection of Cores

Sites DWA18 (near Cox River) and DWA 35 (near Nattai River) were chosen because both are riverine zones, which is most influenced by the river feeding the reservoir and is characterised by complex sedimentation. The DWA2 come under lacustrine zone (near damwall), where, with increased water depth and slower currents, the water body more closely resembles a lake than a river, characterised by more steady and constant sedimentation [Smol, 2002]. All details of sampling and preservation techniques and experiments performed are described in Chapter II.

5.6 Results and Discussion

5.6.1 Core 1 (near damwall)

The activities of 210Po and 226Ra in sediment core 1 intervals were determined as shown in Table 5.1. 210Po activity, which represents the total 210Pb activity, was plotted against depth (Fig. D1A). 210Po activity decreases exponentially with depth down to 40cm and then slight deviation. The average 226Ra line plotted on the 210Po chart indicates that the bottom two points of the profile may have reached the lake sediments background levels. 226Ra activity, which represents supported 210Pb fraction was plotted against depth (Fig. D1B). The radium graph shows an approximate vertical trend against depth (average value =51 Bq/kg).

Excess 210Pb or unsupported Pb was calculated by subtracting 226Ra activity (a proxy for supported 210Pb) from 210Po activity (a proxy for total 210Pb) for each sediment interval as shown in Table 5.1. Excess 210Pb activity was plotted on a log linear graph against depth (Fig D1C). The excess 210Pb was low after 25 cm probably due to climatic and geographic conditions at the site. Core 1 demonstrates a decay profile from 5 to 25 cm and have a correlation coefficient r2= 0.95. The sedimentation rate for core 1 has been calculated to be 0.47 ± 0.07 (cm/year) using a modified CIC model as described by Brugam (1978).

82 5.6.2 Core 2 (near Cox river)

210Po activity decreases exponentially with depth down to 15 cm (Fig D2 E). 226Ra activity was close to being constant throughout the core, which indicates the sediment, is of the same type and/or source (Fig D2 F). Below 15 cm excess 210Pb activity does not show a decay pattern (Fig D2 G). The calculated sedimentation rate for core 2 is calculated using the excess 210Pb activity from the top 15 cm of the core only. The Pb-210 profile was also normalised using <63 μm size data (Fig D2 I). The calculated sedimentation rate did not change very much. Normalising using < 2μm grain size data gave a poor linear regression result. The calculated sedimentation rate 2 near Cox River is 0.19±0.004 cm/year (r =0.99)

5.6.3 Core 3 (near Nattai river)

The activities of 210Po and 226Ra in sediment core 3 intervals from Lake Burragorang were determined as shown in Table 5.3. 210Po activity decreases with depth down to 25 cm as shown in Fig D3 J. 226 Ra activity was close to being constant throughout the core, which indicates the sediment is of the identical type / source (Fig D3 K). Core 3 (Fig D3 L) shows a decay profile of excess 210Pb up to a depth of 25 cm, from which a sedimentation rate of 0.43±. 0.09 cm/year (r2=0.91) was calculated. The depth of each sediment slice and the corresponding calculated t-values (age) were plotted as shown in Fig. D1 D, D2 H and D3 M.

The ages calculated were used to establish a last 50 years geochronology of changes in organic matter, carbonate content, nutrients and metal concentrations. Effect of climate variability on sediment deposition is widely studied and play major role in determining its composition and deposition [Colliton, 2001; Agnew, 2002; Harrison et al., 2003]. The annual rainfall data at Wallacia Post Office (station no.67029) closest data station to Warragamba Dam (supplied by Bureau of Meteorology, Australia) is plotted against age (Figs 5.2-5.4) and compared with studied parameters mentioned above.

Climate fluctuations connected with the climate phenomenon is called the Southern Oscillation Index (SOI). The SOI is calculated from the monthly or seasonal fluctuations in the air pressure difference between Tahiti and Darwin.

83 Periods of strong protracted negative values relate to El Nin˜o episodes while periods of strong protracted positive values relate to La Nin˜a episodes. Most El Nin˜o events are associated with drought over eastern Australia while La Nin˜a events are associated with above-average rainfall and flooding [Power, 2000]. The most important episodes of La Niña or high rainfall occurred between 1945-1956, 1961 - 1969, 1974, 1976, 1978, 1984, 1988-1990 and a moderate La Niña event occurred in 1998/99, which weakened back to neutral conditions before reforming for a shorter period in 1999/2000 causing widespread flooding throughout Australia [Bureau of Meteorology, 2006].

Correlation was made unto 25 cm depth in core 1 and 3, and 15 cm depth in core 2 as cores demonstrate a decay profile unto these depths only. In core 1 the increase in all metals and nutrients is observed after highest rainfall in 1950 and concentrations continue to increase during elevated period of rainfall from 1961 to 1969. Besides, rainfall correlation this exalted period in metal concentrations is also coincident with the construction of Warragamba dam during the period between 1948 to1962. Co, Cu, Fe, Mn Pb and V showed fluctuations around 1987, coincident with La Niña event, which occurred around this time (1978-1988). A steady increase was observed in percentage of organic matter and carbonate and no correlation was found with rainfall events. No correlation was found between rainfall and organic matter content and carbonate.

Metals and nutrients at Core 3 also showed strong correlation with 4 powerful La Niña events occurred during the period from 1950 to 1978. The metals thereafter continuously decreased until around 17 years ago, when they begin to increase again. This increase correlates with the moderate rainfall around 1990 and two hazard reduction bushfires around this area followed by rain in 1998.

The increase in percentage of organic matter and carbonate content between 25 and 17 years ago could be attributed to post fire rainfalls after the bushfires occurred during the period between 1981-1985. Sedimentation at Core 2 is low compared to other two locations and small segment (0-15cm) of it showed decay profile, however, metals and nutrients values displayed good correlation with La Niña episodes.

84 No strong explanation could be given for decrease in concentrations but it may be argued that since lake level is going down, wind induced turbulence may enhance the sediment resuspension and release pollutants in water column from sediments.

85 Table 5.1. Activity variation of 210 Po, 226Ra and excess 210Pb with depth in sediment core 1

Core 1 Depth (cm) Activity of Activity of Activity of from-to Po-210 Ra-226 excess Pb-210 (Bq/kg) (Bq/kg) (Bq/kg) 1 0 - 5 105 +/- 2 56+/- 3 50+/-4 2 5-10 93 +/- 2 66+/- 4 27+/-5 3 10-15 70 +/-2 51+/- 3 19+/-3 4 15- 20 64 +/-2 47+/- 3 17+/-4 5 20 - 25 56 +/-1 44+/- 3 12+/-3 6 30 - 35 53 +/-1 57+/- 3 NA 7 35 - 40 38 +/-1 42+/- 3 NA 8 40 - 45 46 +/-2 41+/- 2 5+/-3

Table 5.2. Activity variation of 210 Po, 226Ra and excess 210Pb with depth in sediment core 2

Core 1 Depth (cm) Activity of Activity of Activity of from-to Po-210 Ra-226 excess Pb-210 (Bq/kg) (Bq/kg) (Bq/kg) 1 0 - 5 95.4 +/- 3 40.6+/- 2.6 54.8+/-4.0 2 5-10 70.5 +/- 1.9 45.4+/- 2.9 25.1+/-3.4 3 10-15 57.0 +/-1.3 46.1+/- 2.8 10.9+/-3.1 4 15- 20 61.7 +/-1.7 42.1+/- 2.6 19.6+/-3.0 5 20 - 25 57.6 +/-2.5 39.2+/- 2.5 18.4+/-3.5 6 30 - 35 51.7 +/-2.1 39.5+/- 2.4 12.3+/-3.2 7 35 - 40 51.8 +/-2.4 38.6+/- 2.4 13.2+/-3.4

Table 5.3. Activity variation of 210 Po, 226Ra and excess 210Pb with depth in sediment core 3

Core 1 Depth (cm) Activity of Activity of Activity of from-to Po-210 Ra-226 excess Pb-210 (Bq/kg) (Bq/kg) (Bq/kg) 1 0 - 5 66.0+/- 2.3 36.8+/- 2.2 29.2+/-3.2 2 5-10 66.2+/- 1.8 36.2+/- 2.4 30.0+/-3 3 10-15 50.4+/-1.8 37.4+/- 2.2 13.0+/-2.9 4 15- 20 50.7+/-1.5 38.7+/- 2.4 12.0+/-2.8 5 20 - 25 37.6+/-1.1 29.9+/- 1.8 7.7+/-2.1 6 30 - 35 39.2+/-1.0 29.2+/- 1.8 10.0+/-2.1 7 35 - 40 30.1+/-1.0 29.2+/- 1.9 0.9+/-2.2

86 Concentration (mg/kg)

500 1000 1500 200010000 2500 20000 3000 30000 40000 50000 0 0 2000 5 1990 10 10 15 1980 20 20 1970 25 1960 Year 1950 Depth (cm) 30 30 35 1940 40 40 1930

1 Mn Fe TP TN

(%) 0 102030405060708090100 0 2000 10 1990 1980 20 1970

1960 Year 1950

Depth (cm) Depth 30 1940 40 1930

% Organic Matter % Carbonate % moisture

Concentration (mg/kg) 0 20 40 60 80 100 120 0 2000 1990 10 1980

20 1970

1960 Year

Depth (cm) 30 1950 1940 40 1930 As Cd Cr Co Cu Pb Ni Se V Zn

2005 1995 1985

1975 ar 1965 Ye 1955 1945 1935 0 500 1000 1500 2000 Rainfall (mm)

Rainfall Bushfire

Fig 5.2. Lake Burragorang Core 1 age versus 1) Rainfall 2) Metals 3) Organic matter and Carbonate content 4) Nutrients, Fe and Mn

87 Concentration (mg/kg)

100 600 1100 160010000 2100 20000 30000 40000 50000 0 0 1990 5 5 1970 10 10 1950 15 15 1930

20 20 1910 Year 25 25

Depth (cm) 1890 30 30 1870 35 35 1850

Mn Fe TP TN

(%) 0 102030405060708090100 0 5 1990 1970 10 1950 15 1930 20 1910 25 1890 Depth (cm) Depth 30 1870 35 1850

% Organic Matter % Carbonate % moisture

Concentration (mg/kg) 0 102030405060708090 0 1990 5 1970 10 1950 15 1930 20 1910 Year

Depth (cm) 25 1890 30 1870 35 1850 As Cd Cr Co Cu Pb Ni V Zn

2005 1995 1985 1975 1965 Year 1955 1945 1935 0 500 1000 1500 2000 Rainfall (mm)

Rainfall Bushfire

Fig 5.3. Lake Burragorang Core 2 age versus 1) Rainfall 2) Metals 3) Organic matter and Carbonate content 4) Nutrients, Fe and Mn

88 Concentration (mg/kg)

300 800 130010000 1800 2300 20000 30000 40000 0 0 2000 5 5 1990 10 10 1980 15 15 1970

20 20 1960 Year 25 25 1950 Depth (cm) 30 30 1940 35 35 1930

Mn Fe TP TN

(%) 0 1020304050607080 0 2000 5 1990 10 1980 15 1970 20 1960 Year 25 1950 Depth (cm) Depth 30 1940 35 1930

% Organic Matter % Carbonate % moisture

Concentration (mg/kg) 0 1020304050607080 0 2000 5 1990 10 1980 15 1970 20 1960 Year

Depth (cm) 25 1950 30 1940 35 1930 As Cd Cr Co Cu Pb Ni V Zn

2005 1995 1985 1975 1965 Year 1955 1945 1935 0 500 1000 1500 2000 Rainfall (mm)

Rainfall Bushfire

Fig 5.4. Lake Burragorang Core 3 age versus 1) Rainfall 2) Metals 3) Organic matter and carbonate content 4) Nutrients, Fe and Mn

89

Chapter VI. Conclusion

The present thesis reports the distribution of As, Cd, Cr, Co, Cu, Fe, Pb, Mn, Hg, Mo, Ni, Se, V and Zn in sediments of Lake Burragorang. In surfacial sediments concentrations of Hg and Se in all locations (except at DWA3 and DWA2) were found below the detection limit (0.1 mg/kg). Sites DWA2, DWA9 and DWA27 appeared to be most polluted sites as almost all metal levels are above the estimated background values, however, DWA19 found to be least polluted. The metal concentration generally decreases in the order Fe >Mn >Zn >V >Cr >Pb ≅Ni ≅Cu>Co >As> Mo>Se> Cd as was reported by Fytianos and Lourantou [2004]. Overall metal distribution picture depicted that locations close to damwall and middle of the lake are more polluted compared to others. This may attribute to proximity of sources. Werri Berri (Monkey Creek) catchment is close to the dam wall (approximately 4 Km from the offtake point for Sydney's water supply) fairly urbanised and the most developed area in the Warragamba Special Area. Water quality problems have been found in the upper part of the catchment including high levels of turbidity, iron, nutrients and faecal bacteria. Cryptosporidium and Giardia have been detected in storm water channels draining from the Oak township to Werri Berri Creek. Oakdale colliery, which ceased operation in 1999, is in high-risk categories [DEC, 2005] located near the identified polluted sites in this study. Based on guidelines given by Long et al [1995] the concentration of Cd, Cr, Hg, Pb and Zn were found below the effects range-low (ERL) whereas Cu levels were close to ERL at M3 and DWA2. Arsenic and Ni were present at higher concentration than ERL at DWA2 and DWA9. Ni also exceeded the ERL at DWA27. Mn was found above ERL at DWA35 and DWA18 and effects range-median (ERM) at DWA3, DWA2, DWA9, DWA12 and DWA35. Interestingly Fe was found to be above ERL at all sites and it is a matter of great concern that it even exceeded the ERM at DWA2, DWA9, DWA27 and M3 which make these stations poor on rating.

This is the first study to report metal speciation data for lake Burragorang sediments. The possible bioavailability of these metals was assessed using sequential extraction. With a few exceptions here and there the speciation profile of a particular metal is 90 same throughout the stretch of Lake Burragorang that has been covered under this study. The speciation patterns of As, Fe, Mo, Ni, Pb and V indicate their significant association with the residual fractions of sediments. Small percentage of Mo is hosted by first two phases mainly at upstream. Cu and Cr speciation demonstrated their high percentage association with residual and organic fraction and make them least mobile. Substantial amount of metals like Cd, Co, Mn, and Zn are present in the first three fractions exchangeable, carbonate and reducible. The exchangeable and carbonate, which are considered to be weakly bound fractions and may equilibrate with the aqueous phase thus becoming more bioavailable. The Fe-Mn oxide and the organic matter have a scavenging affect and may provide a sink for heavy metals. The release of the metals from this matrix will most likely be affected by the redox potential and pH. Moderate association of Ni and Pb in carbonate fractions and Fe- Mn oxide fractions thus has a possibility of becoming readily bioavailable. The total Fe in the sediments is quite high and even its lower amount bound to the exchangeable and carbonate fraction could cause deleterious effects. Overall, data on the fractional distribution of heavy metals indicate that Cd, Co, Mn, and Zn have the highest migration mobility whereas Cu and Cr least in Lake Burragorang sediments. The results showed that leaching of metals from sediments from highest to lowest is in the following order: Mn=Cd>Co= Zn > Ni > Mo> Pb>Fe>V> As>Cu>Cr.

Sediments cores were also analysed as they provide a historical record of the various influences on the aquatic system by indicating both natural background levels and the man-induced accumulation of elements over an extended period of time and can be used to know the spatial distribution of heavy metals in sediment depth profile. Cores were investigated for carbonate content, organic matter, nutrients and heavy metals. Carbonate content were found more or less constant at all location except DWA35 whereas organic matter decreased with depth on those sites, which are near to dam wall (DWA2, DWA6 and MC3). The background concentration of 14 metals were established by interpreting their concentrations in sediment cores and background values were found as 4.7, 0.2, 23, 12, 20, 29000, 22, 660, < 0.2, 0.25, 19.7, 0.13, 37 and 68 mg/kg for As, Cd, Cr, Co, Cu, Fe, Pb, Mn, Hg, Mo, Ni, Se, V and Zn, respectively. The background levels are quite comparable to other studies.

Total phosphorus concentrations at Lake Burragorang were found higher than Bellinger Estuary (TP 176 mg/Kg) in northern New South Wales, which is 91 considered to be almost pristine [Birch et al., 1999]. Nutrient concentrations in the bottom sediments varied substantially among the different sites. Most of them showed positive trend (that is, nutrient concentration increased toward the top of the sediment core). These trends in nutrient concentrations may be related to an increase in fertilizer use, livestock production and sewage-treatment plants around the catchment. Alternatively, the trends may be indicative of diagenesis (that is, post depositional changes in the sediment caused by various processes including decomposition) [Juracek, 2004].

In view of usual anoxic conditions below the surfacial sediment and where sulphide is prominent phase to control the bioavailability, the Acid volatile sulphide (AVS) and simultaneously extracted metal (SEM) method was used to predict the availability of selected heavy metals (Cd, Cu, Ni, Pb and Zn) for different organisms on selected sites.

The results showed that these simultaneously extracted metals at all stations were higher than AVS and their ratio was found greater than 1, which indicates that available AVS is not sufficient to bind with the extracted metals. On this basis it can be concluded that AVS is not a major metal binding component for Lake Burragorang sediments and contained metals potentially bioavailable to benthic organisms. AVS concentration depends on season and depth. Low concentration of AVS occurred in the winter because in colder temperatures, FeS formation rates were lower and a smaller supply of FeS-rich particles was brought up from below by bioturbation [Aller, 1977].

Sediments, collected for the present study for SEM and AVS analysis were in winter season and hence the low levels of AVS were found. Low AVS concentrations indicate that most metals are bound by sediment constituents other than AVS [Allen et al., 1993]. A study conducted by Lawrence Berkeley National Laboratory (LBNL) also had sites with SEM-AVS values greater than one due to relatively low AVS values and not necessarily high concentrations of metals. No toxicity to benthic organisms was observed from these LBNL sites [2001]. Other constituents in the sediment, such as iron and manganese oxides and organic matter, may have decreased the bioavailability of heavy metals [Di Toro et al., 1990]. The unbound metals toxicity to benthic organisms can be explained by analyzing individual SEM 92 concentrations according to their upper effects threshold (UET) levels (Table 4.7). All the locations had individual SEM concentrations lower than their UET. Even though these investigated metals were bioavailable in the sediment, their individual metal concentrations are not expected to be toxic to benthic organisms.

Cores were also subjected to 210Pb dating to determine rate of sedimentation to interpret the depositional history and health of lake environments. Constant initial concentration (CIC) model has been applied in this study to determine sedimentation rates and age profiles. The accumulation of sediment near damwall (0.47 ± 0.07 (cm/year) and near Nattai River inflow (.43±. 0.09 cm/year) is more or less same.

However, near Cox river, sedimentation rate (0.19±0.004 cm/year) is low compared to other two locations.

The ages calculated were used to establish last 50 years geochronology of changes in organic matters, carbonate contents, nutrients and metal concentrations.

In core 1 the increase in all metals and nutrients is observed after highest rainfall in 1950 and concentrations continue to increase during elevated period of rainfall from 1961 to 1969. Besides, rainfall correlation this exalted period in metal concentrations is also coincident with the construction of Warragamba dam during the period between 1948 to1962. As, Co, Cu, Fe, Mn Pb and V showed fluctuations around 1987, which is coincident with La Niña event that occurred approximate this time (1978-1988). A steady increase was observed in percentage of organic matter and carbonate and no correlation was found with rainfall events.

Metals and nutrients at Core 3 also showed strong correlation with 4 powerful La Niña events, which occurred during the period from 1950 to 1978. The metals then continuously decreased until around 17 years ago, and begin to increase after that. This increase correlates with the moderate rainfall around 1990 and two hazard reduction bushfires around this area followed by rain in 1998.

The increase in organic matter and carbonate contents between 25 and 17 years ago could be attributed to post fire rainfalls after the bushfires occurred during the period between 1981-1985. Sedimentation at Core 2 is low compared to other two locations and small segment (0-15cm) of it showed decay profile, however, metals and nutrients values displayed good correlation with La Niña episodes. 93 No strong explanation could be given for decrease in concentrations but it may be argued that since lake level is going down, wind induced turbulence may enhance the sediment resuspension and release pollutants in water column from sediments.

The chemical data collected from Lake Burragorang sediments for metal concentrations and their bioavailability provided information with reasonable approximation. The exact bioavailability is not only influenced by metal geochemistry in sediments, but is also dependent on the physiology and biochemistry of the benthic invertebrates. Further work is needed to understand the degree of bioavailability of these metals, bound with different geochemical phases of sediments, to benthic organism.

The other factors, which influence bioavailability are redox conditions, seasonal variations, analysis of heavy metal concentrations in interstitial water and overlying water on sediment surface. All of these factors could be useful in future studies to determine metal bioavailability significantly.

The suggested further work will help to evaluate sediment quality guidelines for Australian sediments.

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113

Appendix A

Statistical analysis

Table A-1 Uncertainty measurements for different studied variables

DWA2 at 35cm depth Carbonate Oganic Variables As Cd Cr Co Cu Pb Ni V Zn Mn Fe TP TN content matter mg/kg % Standard deviations 0.1 0 0.5 0.4 0.5 0.6 0.5 0.5 4.1 42 3119 31.7 26.4 0.38 1

Coefficiect of variation 1.8 1.9 2 2.8 1.9 2 2.2 1.3 3.4 8.4 6 3.8 4.4 8.1 5

DWA30 at 30cm depth Carbonate Oganic Variables As Cd Cr Co Cu Pb Ni V Zn Mn Fe TP TN content matter mg/kg % Standard deviations 0.5 0 2.1 1.1 1.5 2.7 2 5.3 6 51 3324 24 36 0.09 0.43

Coefficiect of variation 7.7 8.5 7.5 9.1 6.7 9.5 9.2 9.1 7.7 9.4 8.6 6.9 3 8.1 8.3

Number of multiple runs=6

114 Appendix B

Concentration of organic matter, carbonate content and nutrients

DWA2 DWA6 Variables Variables

0 5 10 15 20 25500 1000 1500 2000 2500 0 5 10 15300 600 900 1200 1500 0 0

10 10

20 20

30 (cm) Depth 30 Depth (cm)

TP (mg/Kg) TP (mg/Kg) 40 40 TN (mg/Kg) TN (mg/Kg) % Organic matter % Organic matter % Carbonate % Carbonate

50 50 DWA15 DWA9 Variables Variables 0 5 10 15350 700 10501400175021002450 0 5 10 15350 700 1050 1400 1750 2100 0 0

10 10

20 20 Depth (cm) Depth (cm) Depth

30 30 TP (mg/Kg) TN (mg/Kg) TP (mg/Kg) % Organic matter TN (mg/Kg) % Carbonate % Organic matter % Carbonate 40 40

Fig B-1. Depth distributions of carbonate content, organic matter and nutrients in sediments

Continued-- 115 Appendix B (continued)

DWA27 UWS15 Variables Variables

0 5 10 15 300 600 900 1200 1500 0 5 10 15 20 500 1000 1500 2000 0 0

5

10 10

15 20 Depth (cm) Depth (cm)

20

TP (mg/Kg) TN (mg/Kg) TP (mg/Kg) 25 % Organic matter 30 TN (mg/Kg) % Carbonate % Organic matter % Carbonate

30 DWA30 UWS14 Variables Variables

0 5 10 15 350 700 1050 1400 1750 2100 0 7 14 21 28 35 600 1200 1800 2400 3000 3600 0 0

5

10 10

15 20 Depth (cm) Depth (cm) TP (mg/Kg) 20 TN (mg/Kg) % Organic matter % Carbonate

TP (mg/Kg) TN (mg/Kg) 25 30 % Organic matter % Carbonate

30

Fig B-2. Depth distributions of carbonate content, organic matter and nutrients in sediments

116

Appendix B (continued)

DWA35 DWA39 Variables Variables

0 5 10 15 400 800 1200 1600 2000 2400 0 5 10 15 20350 700 1050 1400 1750 0 0

10 5

20 10 Depth (cm) Depth (cm)Depth

15 30 TP (mg/Kg) TN (mg/Kg) TP (mg/Kg) % Organic matter TN (mg/Kg) % Carbonate % Organic matter % Carbonate

40 20 UWS13 DWA12 Variables Variables

0 5 10 15 200 400 600 800 1000120014001600 0 5 10 15 20 500 1000 1500 2000 2500 3000 3500 0 0

10 10

20 20 Depth (cm) Depth (cm)

TP (mg/Kg) TN (mg/Kg) TP (mg/Kg) 30 TN (mg/Kg) % Organic matter % Organic matter % Carbonate 30 % Carbonate

40

Fig B-3. Depth distributions of carbonate content, organic matter and nutrients in sediments

117

Appendix B

DWA18 MC3 Variables Variables

0 5 10 15 300 600 900 1200 1500 1800 2100 2400 0 5 10 15 20 500 1000 1500 2000 2500 0 0

5 10

10

20 Depth (cm) Depth (cm) Depth 15

30 20 TP (mg/Kg) TP (mg/Kg) TN (mg/Kg) TN (mg/Kg) % Organic matter % Organic matter % Carbonate % Carbonate 40 25

Fig B-4. Depth distributions of carbonate content, organic matter and nutrients in sediments

118 Appendix C

Concentrations of metals

Metal (mg/Kg) Metal (mg/Kg)

0 20406080100120140 500 1000 1500 30000 40000 50000 0 0

10 10

20 20

30 30 Depth (cm) Depth (cm) Depth

40 40

50 50 DWA2 Metal (mg/Kg) Metal (mg/Kg)

0 20 40 60 80 100 120 140 160 180 600 1200 1800 32000 40000 0 0

10 10

20 20

30 30 Depth (cm) Depth (cm)

40 40

50 50

Metal (mg/Kg) DWA6 Metal (mg/Kg)

0 20406080100120140 600 1200 1800 2400 3000 30000 40000 0 0

10 10

20 20 Depth (cm) Depth (cm)

30 30

DWA9

As Co Hg Se Fe Cd Cu Mo V Mn Cr Pb Ni Zn Fig C-1. Depth profiles of metals in sediments

119 Appendix C (continued)

Metal (mg/Kg) Metal (mg/Kg) 800 1600 2400 27500 33000 38500 0 20406080100120 0 0

10 10

20 20 Depth (cm) Depth Depth (cm) 30 30

40 40 DWA15 Metal (mg/Kg) Metal (mg/Kg)

0 20 40 60 80 100 120 140 800 1600 2400 36000420004800054000 0 0

5 5

10 10

15 15 Depth (cm) Depth Depth (cm) Depth 20 20

25 25

30 30 DWA27

Metal (mg/Kg) Metal (mg/Kg)

0 20406080100120 600 1200 1800 35000 40000 45000 50000 0 0

10 10

20 20 Depth (cm) Depth Depth (cm) Depth

30 30

As Co Hg Se Fe UWS15 Cd Cu Mo V Mn Cr Pb Ni Zn Fig C-2. Depth profiles of metals in sediments

120 Appendix C (continued)

Metal (mg/Kg) Metal (mg/Kg) Metal (mg/Kg) Metal (mg/Kg) 0 20406080 350400450500550 28000 32000 36000 800 1600 2400 27500 33000 38500 0 0 20406080100120 0 0 0

10 10 10 10

20 20 20 20 Depth (cm)Depth Depth (cm) Depth Depth (cm) Depth Depth (cm) 30 30 30 30

40 40 40 DWA30 DWA15 Metal (mg/Kg) Metal (mg/Kg) Metal (mg/Kg) Metal (mg/Kg) 0 20406080100 200 250 300 350 33000345003600037500 0 0 20 40 60 80 100 120 140 0 800 1600 2400 36000420004800054000 0 0

5 5 5 5

10 10 10 10 15 15

15 Depth (cm) 15 Depth (cm) Depth 20 20 Depth (cm) Depth Depth (cm) Depth 20 20 25 25

25 25 30 30 UWS14 30 30 Metal (mg/Kg) DWA27 Metal (mg/Kg)

0 204Metal (mg/Kg)06080 400 600 800Metal 30000 (mg/Kg) 35000 40000 0 0 0 20406080100120 600 1200 1800 35000 40000 45000 50000 0 0 10 10

10 2010 20 Depth (cm) Depth Depth (cm) Depth 30 20 20

30 (cm) Depth Depth (cm) Depth

40 40 30 30 DWA35

As Co Hg Se Fe As Co Hg Se Fe UWS15 Cd Cu V Mn Cd Cu Mo MoV Mn Pb Cr PbCr Ni ZnNi Zn Fig C-3. Depth profiles of metals in sediments

121 Appendix C (continued)

0 20406080 300 330 360 26000 28000 30000 0 0

5 5

10 10 Depth (cm) Depth (cm)

15 15

20 20 DWA39

Metal (mg/Kg) Metal (mg/Kg)

0 20406080 200 300 400 20000225002500027500 0 0

10 10

20 20 Depth (cm) Depth (cm)

30

30

40

UWS13 Metal (mg/Kg) Metal (mg/Kg)

0 20 40 60 80 100 120 140 800 160024003200 27000315003600040500 0 0

10 10

20 20

Depth (cm) 30 Depth (cm) 30

40

40 DWA12 As Co Hg Se Fe Cd Cu Mo V Mn Cr Pb Ni Zn

FigC-4. Depth profiles of metals in sediments 122 Appendix C

Metal (mg/Kg) Metal (mg/Kg)

0 20406080 300 450 600 750 2700030000330003600039000 0 0

10 10

20 20 Depth (cm) Depth Depth (cm) Depth 30 30

40

40

DWA18

Metal (mg/Kg) Metal (mg/Kg)

0 20406080100 1200 2400 3600 42500 51000 59500 68000 0 0

5 5

10 10 Depth (cm) Depth Depth (cm) Depth 15 15

20 20

25 25

MC3

As Co Hg Se Fe Cd Cu Mo V Mn Cr Pb Ni Zn

Fig C-5. Depth profiles of metals in sediments

123 Appendix D Sedimentation rate

A B

Po-210 Activity (Bq/kg) Ra-226 Activity (Bq/kg)

0.00 50.00 100.00 150.00 0.00 50.00 100.00 0 0

10 10 20 20 30 30 40 Depth (cm) Depth (cm) 40 50

50 Vertical line = average Ra- 60 Vertical line = average Ra-226 226 60 70

C D

Pb-210 (excess) Activity (Bq/kg) Age (years) 1.00 10.00 100.00 0204060 0 0

10

20 10

30 Depth (cm) Depth (cm) 20 40

50

30 60

Fig D-1. Core 1 profile of A) Po210 B) Ra210 C) excess Pb210 activity and D) age versus depth

124 Appendix D (continued)

E F

Po-210 Activity (Bq/kg) Ra-226 Activity (Bq/kg) 0.00 50.00 100.00 150.00 0.00 50.00 100.00 150.00 0 0

10

10 20

30 Depth (cm) Depth (cm) 20

40 Vertical line = average Ra-226 Vertical line = average Ra-226

50 30

G H

Pb-210 (excess) Activity (Bq/kg) Age (years) 1.00 10.00 100.00 0 20406080100 0 0

10

20 10

30 Depth (cm) Depth (cm)

40

20 50

I

Pb-210 (excess) Activity (Bq/kg) (normalised with <63 µm grain size) 1.00 10.00 100.00 0

10

Depth (cm) 20

30

Fig D-2. Core 2 profile of E) Po210 F) Ra210 G) excess Pb210 activity and H) age I) excess Pb210 activity normalised with <63 μm size versus depth

125

Appendix D

J K

Po-210 Activity (Bq/kg) Ra-226 Activity (Bq/kg) 0.00 20.00 40.00 60.00 80.00 0.00 20.00 40.00 60.00 80.00 0 0

10 10

20 20

30 30 Depth (cm) Depth (cm)

40 40

Vertical line = average Ra-226 Vertical line = average Ra- 50 50 226

L M

Pb-210 (excess) Activity (Bq/kg) Age (years) 1.00 10.00 100.00 0 20406080 0 0

10

10 20 Depth (cm)

Depth (cm) 30 20

40

30 50

Fig D-3. Core 1 profile of A) Po210 B) Ra210 C) excess Pb210 activity and D) age versus depth

126