Distribution and Speciation of Heavy Metals in Sediments from … · 2019-06-24 · Sharron, Sue, Burhan, Maree and Gavin for the moral support and encouragement. My thanks are due

1

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

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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.

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)

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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

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4.1 INTRODUCTION................................................................61

4.2 STUDY AREA ...................................................................63


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 210

PB 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


5.6.1 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

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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, 226

Ra and excess 210

Pb with depth in

sediment core 1 ................................................................................................. 86


Po, 226

Ra and excess 210

Pb with depth in

sediment core 2 ................................................................................................. 86


Po, 226

Ra and excess 210

Pb with depth in

sediment core 3 ................................................................................................. 86

Table A-1 Uncertainty measurements for different studied variables .......... 114

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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 4.1. AVS and SEM distribution with depth ................................................... 77

Fig 5.1. Pathways by which 210

Pb 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


matter and Carbonate content 4) Nutrients, Fe and Mn ............................. 88


matter and carbonate content 4) Nutrients, Fe and Mn............................... 89

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

nutrients in sediments.................................................................................... 115







Fig C-1. Depth profiles of metals in sediments................................................. 119


Fig C-3. Depth profiles of metals in sediments ................................................ 121

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


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

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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

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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

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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 210

Pb 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 210

Pb 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.

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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

13

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

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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

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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%

16

described the Coxs catchment and its activities in his Masters thesis.

Fig-1.1.

of the water that is stored in Lake Burragorang. Siaka [1998] comprehensively

Warragamba catchment showing Lake Burragorang [SCA, 1999]

17

Table 1.1. Warragamba catchment and its activities

everal rivers join Coxs River before it enters Lake Burragorang. Main tributaries

clude Kowmung, Jenolan and Kedumba Rivers in the lower catchment. Pipers

nt and

they provide New South Wales (NSW) with approximately 30% of its coal-

Warragamba

subcatchment

Area

(Sq

Km)

Total

area

(%)

Inflow

(%)

Major

urban

areas

Major activities

contributing to

pollution

Coxs 2630 29 30

Lithgow,

South

Katoomba

Power station, STP, Coal

mines, Other small

industries -Copper ore

Refining, Pottery, Brick

and Pipe works

Nattai 369.1 4 11.5

Mittagong,

Bowral,

Mossvale

STP, Ceased coalmines,

Swimming pool,

Discharge from industrial

and urban runoff

Wollondilly 3403 37.6 41Goulburn,

Marulan

Agriculture, Grazing,

STP, Pig and Poultry

enterprises,Stables, Meat

and wool processing

Werri Berri 160 2 0.5

Oaks,

Oakdale,W

allacia

Unsewered residential

development,

Agriculture, Livestocks,

Vegetable growing and

Poultry farming

S

in

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 catchme

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

18

the Coxs River and has high levels of

nutrients and suspended solids. Elevated levels of algae are found in Lake

Ncubeeks Creek carries Mn and Al, while Sawyers Swamp Creek carries Se, Mn,

pacts from coal

mine operations are acid mine drainage, salinity and sedimentation. Most coals

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

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 –

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 im

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

19

catchment [Maidment, 1991]. Swaine [1990] had analysed coal samples and found

any

commercial services and supports a range of light industries. Heavy metals in

AWT, 2001]. The catchment covers an area of 369.1 Km in the southeast sector of

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

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 m

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;

2

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].

20

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].

21

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

22

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

23

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

24

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].

25

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 210

Pb 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.

26

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 210

Pb

dating method as described by Brugam [1978] and thus the variation in metals and

27

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

28

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

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

100dim

dimdimx

weightentsewet

weightentsedryweightentsewetcontentMoisture

−=

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.

100dim

% xweightentseinitial

weightcruciblefinalweightcrucibleinitialmatterOrganic

−=

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

K2SO4 and copper sulphate (CuSO4.5H2O) at 390 oC 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

35

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

36

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

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

STIR 5 Hr RESIDUE + 1M NaOAC (10 ml)

pH 5.0, 25 ± 2 ºC

HEAT, 100 ºC, RESIDUE + 0.04 M NH2OH.HCl

6 hr IN 25% CH3COOH (20 ml)

RESIDUE + 0.02 M HNO3 (5ml) + 30%

H2O2 (5 ml), pH 2.0, 100 ºC, 2 hr; 5 ml

30% H2O2 pH 2.0, 100 ºC, 5 hr; 3.2 M

NH4OAC IN 20% (v/v) HNO3,

CONTINUOUS STIRRING, 30 min

DIGESTION RESIDUE + 10 ml REVERSE AQUA REGIA,

45 min, 60 °C on ULTRASONIC BATH, THEN on HOT PLATE

for 45 min at 145 °C

1g SEDIMENT

CENTRIFUGE

CENTRIFUGE

CENTRIFUGE

SUPERNATENT

(CARBONATE

FRACTION)

SUPERNATENT

(Fe-Mn OXIDE

FRACTION)

CENTRIFUGE

RESIDUAL FRACTION SUPERNATENT

(ORGANIC FRACTION)

SUPERNATENT

(EXCHANGABLE

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

39

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 210

Pb analysis was first suggested in 1963 by Goldberg

[1963] and was first applied to lake sediments by Krishnaswamy et al. [1971]. The

total 210

Pb activity was determined by measuring its granddaughter 210

Po, which was

assumed to be in secular equilibrium with 210

Pb. Supported 210

Pb was approximated

by measuring 226

Ra activity.

Approximately 2g of each sample (dry weight) were spiked with 209

Po and 133

Ba

yield tracers to determine the chemical recovery of 210

Po and 226

Ra 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 210

Pb activity whereas 210

Po

activity is in equilibrium with total 210

Pb activity. Calculation of the unsupported

210Pb values was carried out by subtraction of the supported

210Pb activity from the

total 210

Pb activity. The sedimentation rates were calculated using the modified

constant initial concentration (CIC) model method described by Brugam [1978] and

using the formula:

⎟⎠⎞⎜⎝

⎛=At

AoIn

ytI *

1

where,

Ao = unsupported 210

Po at the sediment surface in decays per minute per gram

(dpmg-1

) of dry sediment;

At = unsupported 210

Po activity at time t in dpmg-1

of dry sediment;

y = decay constant of 210

Po (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 210

Po 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

Observed

values

Certified

values

Observed

values

Certified

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

Metal

(mg/kg)

AGAL-10 AGAL-12

n =5 *= Standard deviation

41

42

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

43

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

44

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.

45

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

46

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

47

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.

48

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

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 (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

Metal contaminants in sediments

*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

oncentration of Hg and Se at all locations (except at DWA3 and DWA2) were

values, however, DWA19 was found to be least polluted.

S.No. Station As Cd Co Cr Cu Fe Hg Mn Mo Ni Pb Se V Zn

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 70Lake

Burragorang 4.7 0.2 12 23 20 28500 <0.1 660 0.25 19.7 22 0.13 37 68

mg/kg

Background

Value

C

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

50

0

5

10

15

20

51

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

Burragorang

rsenic, Cu, Mo and Zn exceeded the background limits at sites DWA18 and M3.

her than

background limits. Samples at sites DW 3 and DWA12 contained Mn about 3-5

mes 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

A

Near Werri Berri at M7 Co, Cu and Zn concentrations were discovered hig

A

ti

sites

Con

cent

ratio

n (m

g/K

g)

As Cd Co Mo

DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7

0

20

40

60

80

100

120

sites

Con

cent

ratio

n (m

g/K

g)

Cr Cu Ni Pb V Zn


sites

0800

1600240032004000

30000

40000

50000

Con

cent

ratio

n (m

g/K

g)

Fe Mn


52

A35

and DWA18 and ERM at DWA3, DWA2, DWA9, DWA12 and DWA35.

rang is outlined in Table3.3 and presented graphically

3.5. Mercury and Se were not considered for speciation due

to their low concentration observed in lake sediments. The fractionation profiles

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 DW

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 Burrago

in Figs 3.2, 3.3, 3.4 and

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.

.

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

DW

2 nd 40.0 24.5 1.3 1.2 7.1 32.0 1.5 13.9 6.2 0.5 16.6

A3 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

54

ts

ee 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

and DWA9, but decreases towards the Werri Berri Creek at M7, where it accoun

for only 14% of the total Other than that Cd is dominated by first thr

Sites


Per

cent

age

(%)

Fra

ctio

n of

As

0

20

40

60

80

100

120

55


sequential extractions

Exchangeble Carbonate Fe-Mn Oxide Organic Inert

Sites


Per

cent

age

(%)

Fra

ctio

n of

Cd

0

20

40

60

80

100

120

Sites


Per

cent

age

(% F

ract

ion

of C

o

60)

0

20

40

80

100

120

56

Per

cent

age

(%)

Fra

ctio

n of

Cr

0

20

40

60

80

100

120


Sites

Per

cent

age

(%)

Fra

ctio

n of

Cu

0

20

40

60

80

100

120


s sequential extraction



Sites

Per

cent

age

(%)

Fra

ctio

n of

Fe

0

20

40

60

80

100

120


Sites

Sites


Per

cent

age

(%)

Fra

ctio

n of

Mo

0

20

40

60

80

100

120


Sites


Per

cent

age

(%)

Fra

ctio

n of

Ni

0

20

40

60

80

100

120

Sites


Per

cent

age

(%)

Fra

ctio

n of

Mn

0

20

40

60

80

100

120

57



.

Per

cent

age

(%)

Fra

ctio

n of

Pb

0

20

40

60

80

100

120

58

ig 3.5.

F Metal distributions in Lake Burragorang sediments determined by


Sites


Per

cent

age

(%)

Fra

ctio

n of

V

0

20

40

60

80

100

120


Sites


Per

cent

age

(%)

Fra

ctio

n of

Zn


Sites

0

20

40

60

80

100

120

59

ochemistry. 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

l., 2003]. The distribution of vanadium is similar to Pb, being mostly dominated by

ert fraction followed by reducible with minor amount in oxidisable. The high

here reducible and exchangeable fractions take over. Zn also appeared in the same

attern as Ni but Zn was also associated with exchangeable portion. The residual

action accounting for 42.8 -48.2% of the total Zn concentration. This result is in

ge

a

in

percentages of Zn total content in residual at all sites except DWA19, M3 and M7

w

p

fr

60

greement with Baruah et al. [1996]; Stone and Droppo [1996]; Ma and Rao [1997];

sediments. With a few exceptions here and there, the speciation profile of a particular

carbonate and reducible. The exchangeable and carbonate,

which are considered to be weakly bound fractions and may equilibrate with the

Overall, data on the fractional distribution of heavy metals indicate that Cd, Co, Mn,

a

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

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,

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.

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.

61

etals is largely controlled by the

ent 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-

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 m

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 treatm

62

ulphide complexes [Di Toro et al., 1990; Allen et al., 1993.; Huerta-Diaz et al.,

1998], thereby controlling the metal bioavailability and subsequent toxicity for

ben rn

(us ed

imultaneously extracted me

nd SEM gives an indication of the potential sediment toxicity.

At ΣSEM/AVS < 1, sediments are interpreted as non-toxic because the reactive

xic

uman

activity such as dredging or prop wash [Shipley et al.].

ulphide as the result of their very low solubility products. Under

anaerobic conditions the mobility of the metals is thus strongly reduced and toxic

s of acid

volatile sulphide and simultaneously extracted metals in Dutch marine and

s

thic biocommunities. The AVS-bound metals, with environmental conce

ually Cd, Cu, Ni, Pb and Zn), are extracted at the same time and are call

tals (SEM). The ratio or the difference between AVS s

a

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 o

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 h

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 s

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 variation

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.

63

surface sediments.

However, AVS could play an important role in binding heavy metals in deep layer

In light of the importance of the AVS- SEM concept for assessing the ecological

Fourteen sediment cores were collected from different sites in Lake Burragorang

A27: It is

located in the middle of lake and (d) DWA35: It is at the inflow site of the Nattai.

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

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.

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

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) DW

Table 4.1. Lake Burragorang monitoring sites

Station name Longitude

(AMG)

Latitude

(AMG)

Depth

(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


4.3.1 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

A39 and UWS13) and Cox river in north

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 (DW

(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

65

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

bservation 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,

ase in concentration with depth.

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

o

DWA35 DWA39 and MC3 showed decre

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.

66

values decrease until 20 cm depth and then rise quickly beyond

20cm. The values at the bottom are bit higher than surface values.

may be indicative of diagenesis (that is, post depositional

changes in the sediment caused by various processes including decomposition)

etal levels before the

extent, if any, of heavy metal contamination can be estimated. Such background

a known pristine region. (c) Direct

measurements of metal concentrations in texturally-equivalent sub-surface core

other hand, TN

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

[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 m

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

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.

Table 4.2.

67

Spatial and vertical distributions of carbonate content, organic

agorang matter and nutrients in sediment cores of Lake Burr

ontinued---

35 645 1590 8.6 2.2 67.3

40 602 1505 2.5 0.6 58.2

17 DWA9 5 680 2150 9.7 2.6 65.2

1518 8.6 2.6 71.2

19 15 563 1542 7.8 2.2 64.7

15 502 1725 7.6 1.5 70.2

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

S.No. Sample ID Depth

(cm)

TP

(mg/kg)

TN

(mg/kg)

Organic

matter (%)

Carbonate

(%)

Moisture

(%)

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

16

18 10 595

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

44

C

68

S.No. Sample ID Depth

(cm)

TP

(mg/kg)

TN

(mg/kg)

Organic

matter (%)

Carbonate

(%)

Moisture

(%)

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

69

f the background approach are that it requires a minimum of field data

nd no quantitative toxicity assessments. The disadvantages are that it may be

ifficult both to find suitable reference sites and to determine what levels are

cceptable "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

f Zn concentrations was observed at 10 cm depth at DWA6, however, beyond that

o significant difference was observed with depth. Concentrations of other metals

oderately decreased with depth, but with a peak at 10 cm. No regular trend was

bserved in Fe profile. At DWA9 metal concentrations except Mn increased to a

Advantages o

a

d

a

o

n

m

o

70

A30

was found similar to DWA 9 profile except a slight escalation in Pb, V and Zn

WA12 Cd, Co, Cu and Ni pattern

are more or less same through out the core. Cr, Mo, Pb and V profile moderately

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,

around middle of lake were

substantially elevated over background levels estimated for lake Burragorang

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 DW

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 D

decrease down

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

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.

Table 4.3. Variation in metal concentrations with depth in sediment

core samples

ontinued---

S.No. Sites Depth

(cm)

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

1 DWA2 5 5.6 0.27 33 13 25 29 nd n

e

d 25 0.76 58 92 1090 381

2 10 6.6 0.30 31 14 27 29 nd 0.13 25 0.69 56 96 1280 487

3 15 6.4 0.28 34 13 23 29 nd n

00

00

d 27 0.74 53 92 980 392

4 20 8.2 0.33 37 17 31 33 nd 0.15 32 0.68 58 120 900 356

5 25 5.8 0.20 31 15 24 24 nd n

00

00

d 27 0.46 46 89 880 345

6 30 6.7 0.20 31

00

17 26 26 nd nd 29 0.45 43 97 820 35000

3 12 21 17 nd n7 35 4.9 0.15 2 d 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

9 DWA6 5 7.3 0.26 30 14 29 26 nd nd 30 0.42 52 180 1820 392

10 10 11 0.33 38 17 32 35 nd n

00

d 32 0.29 66 110 1380 413

11 15 10 0.31 39 17 30 34 nd 0.27 33 nd 63 110 1290 432

12 20 7.7 0.33 33 14 24 29 nd 0.14 27 nd 53 93 1210 421

13 25 8.3 0.29 33 15 26 28 nd 0.23 27 nd 53 95 1100 424

14 30 7.0 0.22 32 15 26 27 nd n

00

00

00

00

d 27 nd 49 94 1020 361

15 35 6.6 0.18 31 16 24 26 nd nd 27 nd 47 86 1160 363

16 40 7.1 0.21 31 16 29 26 nd n

00

00

d 28 0.12 42 94 940 365

17 DWA9 5 8.4 0.30 31 17 26 28 nd nd 26 nd 53 96 2860 404

18 10 8.4 0.33 32 15 26 28 nd 0.44 28 nd 52 100 1350 411

19 15 10 0.30 40 20 31 33 nd 0.65 35 nd 66 120 1140 426

20 20 6.5 0.23 36 17 26 29 nd nd 28 nd 52 92 1010 349

21 25 6.3 0.22 27 12 21 25 nd 0.49 21 nd 40 69 730 261

22 30 5.3 0.22 24 13 20 24 nd 0.59 20 nd 36 68 600 255

23 DWA15 5 8.3 0.28 32 16 26 28 nd 4.0 26 nd 55 91 2520 403

24 10 7.8 0.24 31 14 24 26 nd 0.44 25 nd 51 82 1180 356

25 15 7.8 0.26 31 14 24 26 nd 0.29 26 nd 50 90 1000 383

26 20 6.6 0.18 30 14 23 25 nd nd 26 nd 45 83 870 346

27 25 6.0 0.16 30 14 23 23 nd 0.14 24 nd 44 78 1060 33

28 30 4.5 0.22 27 14 22 21 nd nd 22 nd 39 73 810 284

29 35 4.6 0.23 24 14 22 21 nd n

00

00

00

00

00

00

00

00

00

00

00

800

00

d 21 nd 32 69 1060 250

30 DWA27 5 11 0.23 32 16 25 28 nd nd 26 nd 54 85 2460 453

31 10 8.1 0.20 31 14 22 27 nd nd 25 nd 51 79 1390 384

32 15 6.8 0.17 29 14 20 26 nd nd 24 nd 46 69 970 339

33 20 11 0.27 53 24 33 44 nd nd 39 nd

34 25 5.4 0.17 29 12 17 26 nd n

00

00

00

00

88 130 2150 54000

d 19 nd 47 61 800 32000

35 UWS15 5 8.7 0.21 37 18 26 35 nd nd 31 nd 69 90 1800 47200

00

00

0

0

36 10 9.8 0.25 33 16 23 30 nd nd 25 nd 56 79 1450 467

37 15 8.2 0.22 34 14 22 28 nd 0.65 25 nd 56 78 920 428

38 20 7.5 0.24 32 15 21 27 nd nd 24 nd 53 72 1380 3800

39 25 5.4 0.16 30 13 19 27 nd nd 22 nd 47 68 700 3190

40 30 4.3 0.24 34 12 20 28 nd nd 23 nd 50 67 600 315

41 DWA30 5 6.4 0.17 29 14 20 26 nd nd 22 nd 48 72 550 369

42 10 5.8 0.17 29 12 19 26 nd nd 21 nd 47 66 480 347

43 15 6.8 0.42 33 15 22 27 nd nd 24 nd 56 82 500 371

44 20 5.3 0.17 29 12 17 25 nd nd 19 nd 50 70 400 2920

45 25 6.1 0.18 25 10 16 30 nd nd 17 nd 54 75 440 280

46 30 5.5 0.30 24 11 18 22 nd n

00

00

00

00

0

00

d 19 nd 49 69 440 28800

Metals (mg/kg)

C

71

S.No. Sites Depth

(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 10 4.5 0.18 25 11 25 23 nd nd 22 nd 41 78 250 32400

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

52 DWA35 5 6.7 0.17 32 15 20 27 nd nd 22 nd 52 75 710 39800

53 10 6.5 0.21 29 13 18 26 nd nd 20 nd 47 66 540 36700

54 15 7.9 0.19 33 15 23 31 nd nd 23 nd 55 77 580 41400

55 20 6.5 0.20 32 14 20 28 nd nd 23 nd 52 72 470 37200

56 25 4.9 0.12 26 12 16 21 nd nd 18 nd 45 60 440 29200

57 30 4.5 0.10 29 12 16 19 nd nd 18 nd 47 59 450 28700

58 35 4.5 0.17 27 11 15 23 nd nd 17 nd 46 54 470 29900

59 DWA39 5 2.8 0.12 24 11 14 17 nd nd 17 nd 41 57 350 26300

60 10 4.6 0.12 29 12 17 20 nd nd 19 nd 49 65 380 29600

61 15 5.5 0.13 29 12 17 22 nd nd 19 nd 52 63 340 29700

62 UWS13 5 2.0 0.10 28 13 17 19 nd nd 19 nd 47 64 330 25200

63 10 2.6 0.52 29 14 18 20 nd nd 21 nd 51 65 250 25500

64 15 3.6 0.10 28 13 17 19 nd nd 20 nd 51 63 250 25800

65 20 2.2 ND 20 9.3 12 14 nd nd 15 nd 36 45 200 19100

66 25 3.2 ND 34 15 18 18 nd nd 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

68 DWA12 5 9.2 0.23 26 15 25 25 nd 0.29 23 nd 47 86 3290 41700

69 10 7.0 0.23 25 12 22 22 nd 0.3 21 nd 39 82 1420 34000

70 15 6.8 0.20 24 13 24 23 nd 0.21 23 nd 38 87 1060 35200

71 20 6.3 0.20 23 15 23 21 nd 0.18 24 nd 34 110 1040 31100

72 25 5.9 0.20 19 14 24 20 nd nd 22 nd 22 65 1020 30800

73 30 5.9 0.23 20 16 27 20 nd nd 24 0.12 23 72 930 26600

74 35 5.8 0.24 20 15 29 18 nd nd 24 0.24 22 72 720 24100

75 DWA18 5 6.1 0.26 24 12 25 21 nd nd 19 nd 39 79 420 35100

76 10 4.9 0.29 21 11 25 21 nd nd 18 nd 40 86 510 38500

77 15 3.8 0.59 18 9.6 28 21 nd nd 13 nd 37 81 380 26000

78 20 4.9 0.40 23 10 26 23 nd nd 17 nd 39 87 380 29600

79 25 4.7 0.35 24 13 24 23 nd nd 21 nd 35 86 520 28200

80 30 3.7 0.30 21 11 25 19 nd nd 20 nd 27 84 410 25100

81 35 4.0 0.29 21 12 23 19 nd nd 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 10 6.8 0.21 23 15 21 28 nd nd 17 nd 45 79 4210 68000

84 15 7.3 0.19 24 15 21 28 nd nd 18 nd 45 74 790 41500

85 20 6.5 0.18 25 14 20 25 nd nd 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

Metals (mg/kg)

72

nd = Not detected

DL = Method detection limit M

73

Table 4.4.

Background metal levels for Lake Burragorang from sedimentary

metal concentrations

Station

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

Metal (mg/kg)

*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

Metals

(mg/kg)

Lake

Burragorang

Hawkesbury

Rivera

Georges

River/Port

Hackingb

Coxs

Riverc

Sydney

Habourb

Crust

abundanced

Sedimentd

Shallow

water

sedimente

Rural

reach

South

Creekf

Shaleg

Others

As 4.7 - - - - - - - - - 3 h

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

Hg <0.2 - - - - - - - - - 0.01-0.24i

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 - - - - -

68 57 43 130 47 75 95 92 65 95

Zn

*F

a [Shotter

e in %. Weight. (-) No establish data

et al., 1995] b [Irvine and Birch, 1998] c [Siaka, 1998] d [Bowen, 1979] e [Wedepohl, 1970] f [Thomas and

Wedepohl, 1961] h [PTI, 1989] i [Syers et al., 1973] Thiel, 1995] g [Turekian and

74

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

WA27. From the above distribution patterns, it appears that the AVS contents

enerally decrease along the profile with peaks at various depths.

all the sites among HCl-extractable metals (SEM) Cd concentrations were lowest

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

ow compared to values reported in the literature for fresh water sediments

season [Aller, 1977; 2001;

rabowski et al., 2001; Machesky et al., 2004; Morse and Rickard, 2004]. Major

ecrease occurred in the winter because in colder temperatures, FeS formation rates

4.3.5 Acid Volatile Sulphide and Simultaneously Extracted Metals

D

g

In

and Zn was highest. Copper, Ni and Pb were intermediate. Vertical distribution of

l

[Machesky et al., 2004]. AVS concentration depends on season and depth. Many

esearchers observed variation in AVS levels with r

G

d

75

able 4.6.

T Concentrations of AVS and SEM alongwith depth in sediments

of Lake Burragorang

D

D

0.47 3.7

DWA27

D

63 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

Depth

(cm)

AVS

μmol/g

Cd

μmol/g

Cu

μmol/g

Pb

μmol/g

Ni

μmol/g

Zn

μmol/g

ΣSEM

μmol/g

SEM/AVS

WA2

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

WA18

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

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

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

WA35

5 0.21 0.0012 0.072 0.068 0.073 0.28 0.49 2.3

10 0.25 0.0012 0.052 0.0

76

Table 4.7. Guidelines for determining metal toxicity to benthic organisms in

freshwater sediments (values in mg/kg) [Grabowski, 2001]

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

eir UET. Even though these investigated metals were bioavailable in the sediment

eir individual metal concentrations are not expected to be toxic to benthic

rganisms.

Metal Threshold effects

level (TEL)

Upper effects

threshold (UET)

Cd 0.596 3

Cu 35.7 86

Pb 35 127

Ni 18 43

Zn 123.1 520

th

th

o

77

Fig 4.1.

AVS and SEM distribution with depth

AVS Profile at DWA2

0

5

10

30

35

40

0.00 0.20 0.40 0.60 0.80

AVS (umol/g)

15

(cm

)

20

25

Dep

th AVS

SEM Profile at DWA18

0

5

10

15

20

25

30

35

0.0 0.1 0.2 0.3 0.4 0.5

SEM (umol/g)

Dep

th (

cm

)

Cd

Cu

Pb

Ni

Zn

AVS Profile at DWA18

0

5

10

15

20

25

30

35

0.00 0.10 0.20 0.30 0.40 0.50 0.60

AVS (umol/g)

Dep

th (

cm

)AVS


0

5

10

15

20

25

30

35

0.0 0.1 0.2 0.3 0.4 0.5

Dep

th (

cm)

SEM (umol/g)

Cd

Cu

Pb

Ni

Zn


AVS (umol/g)

0

5

10

15

20

25

30

35

0.00 0.10 0.20 0.30 0.40

Dep

th (

cm)

AVS


0

5

10

15

20

25

30

35

40

0.0 0.1 0.2 0.3 0.4

SEM (umol/g)

Dep

th (

cm)

Cd

Cu

Pb

Ni

Zn


0

5

10

15

20

25

30

35

40

0.00 0.10 0.20 0.30

AVS (umol/g)

Dep

th (

cm)

AVS

SEM Profile at DWA2

0

5

25

30

35

40

0.0 0.1 0.2 0.3 0.4

SEM (umol/g)

Dep

t

10

15

20h (

cm)

Cd

Cu

Pb

Ni

Zn

78

Chapter V. Sedimentary record of heavy metal pollution of

Lake Burragorang using 210

Pb 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].

79

.2 Lead –210 Radiometric Dating

specially useful here in Australia as it can determine environmental impact since

uropean settlement [Harrison, 2003]. The development of this technique was first

erg [1963], and it was first applied to the dating of lake sediments

by Krishnaswamy et al. [1971]. 210

Pb has been developed for a number of

003]. The 222

Rn in the atmosphere then

decays to 210

Pb, which attaches to airborn particulate matter and either – (A) falls

irectly into the catchment or (B) falls in the catchment region and is washed in by

5

One of the most promising methods of dating on a time scale of 100-200 years is

by means of 210

Pb, a natural radioisotope with a half-life of 22.26 years. This is

e

E

initiated by Goldb

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 226

Ra (half life 1622 yrs) to the inert gas 222

Rn (half life

3.83 days) followed by a series of short lived isotopes to 210

Pb [Brenner et al., 1994].

The 210

Pb 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 210

Pb. The 210

Pb present in sediments is described and

analysed as two components, 'supported' 210

Pb and 'unsupported' 210

Pb (Fig 5.1).

226Ra in the sediment within the catchments area enters via erosion wherein it decays

to 210

Pb. The 210

Pb formed by the 'in situ' decay of 226

Ra 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 210

Pb from other sources. Three components are identified by which excess 210

Pb

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, 2

d

80

222

Rn escapes up the water column due to the

decay of 226

Ra already in the river bed. Part of the 222

Rn will migrate up to the

rain or erosion. In the third route (C)

surface of the water and escape to the atmosphere, where it will decay to 210

Pb 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 210

Pb. 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 210

Pb reaches lake sediments [Oldfield, 1981;

Organo, 2000]

Unfortunately, the activity of 210

Pb cannot be measured directly as it is a beta emitter

and peaks on the spectrum is difficult to distinguish due to substantial background.

Instead, 210

Ra and 210

Po are analysed, as they are alpha emitters. Alpha emitters tend

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 210

Pb. The unsupported 210

Pb activity is determined

from it's granddaughter isotope Polonium-210, which is assumed to be in secular

to

81

g technique can be applied to determine sedimentation rates and age

iles through the use of modeling. There are two main models used for age

determinations using the 210

Pb dating method [Organo, 2000]. The first of these is the

constant rate of 210

Pb supply or CRS model, which assumes that the supply of 210

Pb

to the accreting material is occurring at a constant rate. In this model the initial

unsupported 210

Pb 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 210

Pb 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 210

Pb, and that each layer of sediment will have the same initial

unsupported 210

Pb concentration [Appleby and Oldfield, 1992]. The CIC model has

een applied within this study, however, the two models yield the very similar results

ton et

Turner and Delorme, 1996] .

The current study has been undertaken to study the variability in metals and nutrients

Fo

equilibrium with 210

Pb [Heijnis et al., 1987; Ivanovich et al., 1992; Ravichandran et

al., 1995]. The activity of the 'unsupported 210

Pb in a given sample is found by

subtracting the activity of 226

Ra from the activity of the 210

Po. The 'unsupported 210

Pb

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 210

Pb datin

prof

b

if the accumulation rates are constant and not too large [Oldfield, 1981; Chan

al., 1983; Appleby, 1993;

concentrations through lead-210 dating within lake Burragorang and compared with

past data of rainfall and bushfire.

5.4 Sampling Locations

urteen 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

82

ar Nattai River) were chosen

because both are riverine zones, which is most influenced by the river feeding the

acting 226

Ra activity (a proxy

cess 210

Pb was low after 25 cm probably due to

climatic and geographic conditions at the site. Core 1 demonstrates a decay profile

a correlation coefficient r2= 0.95. The sedimentation rate

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 (ne

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.


5.6.1 Core 1 (near damwall)

The activities of 210

Po and 226

Ra in sediment core 1 intervals were determined as

shown in Table 5.1. 210

Po activity, which represents the total 210

Pb activity, was

plotted against depth (Fig. D1A). 210

Po activity decreases exponentially with depth

down to 40cm and then slight deviation. The average 226

Ra line plotted on the 210

Po

chart indicates that the bottom two points of the profile may have reached the lake

sediments background levels. 226

Ra activity, which represents supported 210

Pb

fraction was plotted against depth (Fig. D1B). The radium graph shows an

approximate vertical trend against depth (average value =51 Bq/kg).

Excess 210

Pb or unsupported Pb was calculated by subtr

for supported 210

Pb) from 210

Po activity (a proxy for total 210

Pb) for each sediment

interval as shown in Table 5.1. Excess 210

Pb activity was plotted on a log linear graph

against depth (Fig D1C). The ex

from 5 to 25 cm and have

for core 1 has been calculated to be 0.47 ± 0.07 (cm/year) using a modified CIC

model as described by Brugam (1978).

83

exponentially with depth down to 15 cm (Fig D2 E). Ra

activity was close to being constant throughout the core, which indicates the

e type and/or source (Fig D2 F). Below 15 cm excess 210

Pb

activity does not show a decay pattern (Fig D2 G). The calculated sedimentation rate

a in sediment core 3 intervals from Lake Burragorang

able 5.3. 210

Po activity decreases with depth down to

25 cm as shown in Fig D3 J. 226

Ra activity was close to being constant throughout

tions. Effect of

climate variability on sediment deposition is widely studied and play major role in

calculated from the monthly or seasonal

fluctuations in the air pressure difference between Tahiti and Darwin.

5.6.2 Core 2 (near Cox river)

210Po activity decreases

226

sediment, is of the sam

for core 2 is calculated using the excess 210

Pb 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

near Cox River is 0.19±0.004 cm/year (r2=0.99)

5.6.3 Core 3 (near Nattai river)

The activities of 210

Po and 226

R

were determined as shown in T

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 210

Pb 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 concentra

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

84

ative values relate to El Ninõ episodes while periods

La Ninã episodes. Most El Ninõ events

are associated with drought over eastern

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

after highest rainfall in 1950 and concentrations

Periods of strong protracted neg

of strong protracted positive values relate to

Australia while La Ninã 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

metals and nutrients is observed

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.

85

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.


Po, 226

Ra and excess 210

Pb with depth in

sediment core 1

86


Po, 226

Ra and excess 210

Pb with depth in

sediment core 2


Po, 226

Ra and excess 210

Pb with depth in

sediment core 3

Po-210

(Bq/kg)

Ra-226

(Bq/kg)

excess Pb-210

(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

Core 1 Depth (cm) Activity of Activity of Activity of

from-to

Core 1 Depth (cm)

from-to

Activity of

Po-210

(Bq/kg)

Activity of

Ra-226

(Bq/kg)

Activity of

excess Pb-210

(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

Core 1 Depth (cm)

from-to

Activity of

Po-210

(Bq/kg)

Activity of

Ra-226

(Bq/kg)

Activity of

excess Pb-210

(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

87


matter and Carbonate content 4) Nutrients, Fe and Mn

1

0

10

20

30

40 19301940

0 10 20 30 40 50 60 70 80 90 100

(%)

Dep

th (

cm)

195019601970198019902000

Yea

r

% Organic Matter % Carbonate % moisture

0

10

20

30

40

0 20 40 60 80 100 120

Concentration (mg/kg)

Dep

th (

cm)

1930

1940

1950

1960

1970

1980

1990

2000

Yea

r

As Cd Cr Co Cu

Pb Ni Se V Zn

05

10152025303540

10000 20000 30000 40000 50000

1930

1940

1950

1960

1970

1980

1990

2000

Yea

r


0

10

20

30

40

2000 2500 3000

Dep

th (

cm)

500 1000 1500

Mn Fe TP TN

1935

1945

1955

1965

1995

2005

0 500 1000 1500 2000

Rainfall (mm)

Ye1975

1985

ar

Rainfall Bushfire

0

5

10

15

20

25

30

35

0 10 20 30 40 50 60 70 80 90 100

(%)

Dep

th (

cm)

18501870189019101930195019701990


0

5

10

15

20

25

30

35

0 10 20 30 40 50 60 70 80 90


Dep

th (

cm)

1850

1870

1890

1910

1930

1950

1970

1990

Yea

r

As Cd Cr Co Cu

Pb Ni V Zn

05

101520253035

10000 20000 30000 40000 50000

18501870189019101930195019701990

Yea

r

Mn Fe TP TN

0

5

10

15

20

25

30

35

100 600 1100 1600 2100


Dep

th (

cm)

1935

1945

1955

1965

1975

1985

1995

2005

0 500 1000 1500 2000

Rainfall (mm)

Yea

r

Rainfall Bushfire

88


matter and Carbonate content 4) Nutrients, Fe and Mn

0

5

10

15

20

25

30

35

0 10 20 30 40 50 60 70 80

(%)

Dep

th (

cm)

19301940195019601970198019902000

Yea

r


0

5

10

15

20

25

30

35

0 10 20 30 40 50 60 70 80


Dep

th (

cm)

1930

1940

1950

1960

1970

1980

1990

2000

Yea

r

As Cd Cr Co Cu

Pb Ni V Zn

05

101520253035

10000 20000 30000 40000

1930

1940

1950

1960

1970

1980

1990

2000

Yea

r

Mn Fe TP TN

0

5

10

15

20

25

30

35

300 800 1300 1800 2300


Dep

th (

cm)

1935

1945

1955

1965

1975

1985

1995

2005

0 500 1000 1500 2000

Rainfall (mm)

Yea

r

Rainfall Bushfire

89


matter and carbonate content 4) Nutrients, Fe and Mn

90

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.

his is the first study to report metal speciation data for lake Burragorang sediments.

With a few exceptions here and there the speciation profile of a particular metal is

T

The possible bioavailability of these metals was assessed using sequential extraction.

91

me throughout the stretch of Lake Burragorang that has been covered under this

study. The speciation patterns of As, Fe, Mo

association with the re ll percentage of Mo is

ainly at upstream. Cu and Cr speciation demonstrated

, Mo, Ni, Se, V

sa

, Ni, Pb and V indicate their significant

sidual fractions of sediments. Sma

hosted by first two phases m

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

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

92

benthic

e

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

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 hav

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

93

in this study to determine sedimentation

rates and age profiles. The accumulation of sediment near damwall (0.47 ± 0.07

d were used to establish last 50 years geochronology of changes in

organic matters, carbonate contents, nutrients and metal concentrations.

was found with rainfall events.

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 210

Pb dating to determine rate of sedimentation to

interpret the depositional history and health of lake environments. Constant initial

concentration (CIC) model has been applied

(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 calculate

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

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.

94

The chemical data collected from Lake Burragorang sediments for metal

The other factors, which influence bioavailability are redox conditions, seasonal

re studies to

determine metal bioavailability significantly.

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.

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.

variations, analysis of heavy metal concentrations in interstitial water and overlying

water on sediment surface. All of these factors could be useful in futu

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

Australian sediments.

95

l and

Mineralogical Forms of Cu and Ni in Contaminated Soils from the Sudbury

2003, "Study of Heavy Metal Pollution and

or the Estimation of Potential

. and Luther, G. W., 1997, "Metals in Surface Waters",

Sleeping Bear Press, Michigan.

llen, H. E., Huang, C. P., Bailey, G. W. and Bowers, A. R., 1995, "Metal

Speciation and Contamination of Soil", Lewis Publishers, London.

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T

Appendix A

Statistical analysis

able A-1 Uncertainty measurements for different studied variables

umber of multiple runs=6

Variables As Cd Cr Co Cu Pb Ni V Zn Mn Fe TP TN

Carbonate

content

Oganic

matter

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

Variables As Cd Cr Co Cu Pb Ni V Zn Mn Fe TP TN

Carbonate

content

Oganic

matter

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

mg/kg %

mg/kg %

DWA2 at 35cm depth

DWA30 at 30cm depth

N

115

Appendix B

Concentration of organic matter, carbonate content and nutrients

Fig B-1.

DWA6Variables

0 5 10 15300 600 900 1200 1500

Dep

th (

cm)

0

DWA2Variables

Depth distributions of carbonate content, organic matter and

nutrients in sediments

ontinued-- C

10

20

30

40

50

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

0 5 10 15 20 25500 1000 1500 2000 2500

Dep

th (

cm)

0

10

20

30

40

50


DWA15DWA9

VariablesVariables

0 5 10 15350 700 1050 1400 1750 2100

Dep

th (

cm)

0

10

20

40

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

0 5 10 15350 700 10501400175021002450

Dep

th (

cm)

0

10

20

30

40


116

Appendix B (continued)

UWS15DWA27

Variables Variables



0 5 10 15 20 500 1000 1500 2000

Dep

th (

cm)

0

10

20

30


0 5 10 15 300 600 900 1200 1500

Dep

th (

cm)

0

5

10

15

20

25

30


DWA30 UWS14Variables Variables

0 5 10 15 350 700 1050 1400 1750 2100

Dep

th (

cm)

0

10

20

30


0 7 14 21 28 35 600 1200 1800 2400 3000 3600

Dep

th (

cm)

0

5

10

15

20

25

30


117

Appendix B (continued)

Fig B-3.

DWA39DWA35

VariablesVariables



0 5 10 15 20350 700 1050 1400 1750

Dep

th (

cm)

0

5

10

15

20


0 5 10 15 400 800 1200 1600 2000 2400

Dep

th (

cm)

0

10

20

30

40


UWS13 DWA12Variables Variables

0 5 10 15 200 400 600 800 1000120014001600

Dep

th (

cm)

0

10

20

30


0 5 10 15 20 500 1000 1500 2000 2500 3000 3500

Dep

th (

cm)

0

10

20

30

40


118

Appendix B

Fig B-4.

DWA18 MC3Variables Variables



0 5 10 15 20 500 1000 1500 2000 2500

Dep

th (

cm)

0

5

10

15

20

25


0 5 10 15 300 600 900 1200 1500 1800 2100 2400

Dep

th (

cm)

0

10

20

30

40


119

endix C

Concentrations of metals

Fig C-1.

App

Depth profiles of metals in sediments

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

DWA2

500 1000 1500 30000 40000 50000

Dep

th (

cm)

0

10

20

30

40

50

0 20 40 60 80 100 120 140

Dep

th (

cm)

0

10

20

30

40

50


DWA6

0 20 40 60 80 100 120 140 160 180

Dep

th (

cm)

0

10

20

30

40

50

600 1200 1800 32000 40000D

epth

(cm

)0

10

20

30

40

50

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

0 20 40 60 80 100 120 140

Dep

th (

cm)

0

10

20

30

600 1200 1800 2400 3000 30000 40

DWA9

000

Dep

th (

cm)

0

10

20

30

FeCo SeHg As Mn Cd Cu VMo

Pb Ni Cr Zn

120

Appendix C (continued)

Fig C-2.



DWA15

0 20 40 60 80 100 120D

epth

(cm

)0

10

20

30

40

800 1600 2400 27500 33000 38500

Dep

th (

cm)

0

10

20

30

40


DWA27

800 1600 2400 36000420004800054000

Dep

th (

cm)

0

5

10

15

20

25

30

0 20 40 60 80 100 120 140

Dep

th (

cm)

0

5

10

15

20

25

30


0 20 40 60 80 100 120

Dep

th (

cm)

0

10

20

30

600 1200 1800 35000 40000 45000 50000

Dep

th (

cm)

0

10

20

30

UWS15FeCo SeHg As Cu Mn Cd VMo Pb Ni Cr Zn


Fig C-3.

121


DWA15

Metal (mg/Kg)

0 20 40 6D

epth

(cm

)0 80 100 120

0

10

20

30

40

Metal (mg/Kg)

800 1600 2400 27500 33000 38500

Dep

th (

cm)

0

10

20

30

40

DWA27

Metal (mg/Kg)

800 1600 2400 36000420004800054000

Dep

th (

cm)

0

5

10

15

20

25

30

Metal (mg/Kg)

0 20 40 60 80 100 120 140

Dep

th (

cm)

0

5

10

15

20

25

30

Metal (mg/Kg)

0 20 40 60 80 100 120

Dep

th (

cm)

0

10

20

30

UWS15

Metal (mg/Kg)

600 1200 1800 35000 40000 45000 50000

Dep

th (

cm)

0

10

20

30

Mn FeAs

Cd Cr

Co Cu

Pb

Hg Mo Ni

SeVZn


DWA30

350400450500550 28000 32000 36000

Dep

th (

cm)

0

10

20

30

0 20 40D

epth

(cm

)60 80

0

10

20

30

40


0 20 40 60 80 100

Dep

th (

cm)

0

5

10

15

20

25

30

200 250 300 350 33000345003600037500

Dep

th (

cm)

0

5

10

15

20

25

30

UWS14


DWA35

400 600 800 30000 35000 40000

Dep

th (

cm)

0

10

20

30

40

0 20 40 60 80

Dep

th (

cm)

0

10

20

30

40

FeCo SeHg As Cu Mn Cd VMo Pb Ni Cr Zn


0 20 40 60 80

Dep

th (

cm)

122

FigC-4. Depth profiles of metals in sediments

0

5

10

15

20

300 330 360 26000 28000 30000

Dep

th (

cm)

0

5

10

15

20

DWA39


UWS13

200 300 400 20000225002500027500

Dep

th (

cm)

0

10

20

30

0 20 40 60 80

Dep

th (

cm)

0

10

20

30

40


0 20 40 60 80 100 120 140

Dep

th (

cm)

0

10

20

30

40

800 160024003200 27000315003600040500

Dep

th (

cm)

0

10

20

30

40

DWA12FeCo SeHg As

Cu Mn Cd VMo Pb Ni Cr Zn

123

Appendix C

Fig C-5.

DWA18

Metal (mg/Kg)

300 450 600 750 2700030000330003600039000

Dep

th (

cm)

0

10

20

30

40

Metal (mg/Kg)

0 20 40 60 80

Dep

th (

cm)

0

10

20

30

40

Metal (mg/Kg)

0 20 40 60 80 100

Dep

th (

cm)

0

5

10

15

20

25

MC3

Metal (mg/Kg)

1200 2400 3600 42500 51000 59500 68000

Dep

th (

cm)

0

5

10

15

20

25

Mn FeAs

Cd Cr

Co Cu

Pb

Hg Mo Ni

SeVZn


Appendix D

124

Sedimentation rate

Fig D-1.

Core 1 profile of A) Po210

B) Ra210

C) excess Pb210

activity and D) age

versus depth

A

0

10

20

30

40

50

60

0.00 50.00 100.00 150.00

Po-210 Activity (Bq/kg)

Dep

th (

cm)

Vertical line = average Ra-226

B

0

10

20

30

40

50

60

70

0.00 50.00 100.00

Ra-226 Activity (Bq/kg)

Dep

th (

cm)


C

0

10

20

30

40

50

60

1.00 10.00 100.00

Pb-210 (excess) Activity (Bq/kg)

Dep

th (

cm)

D

0

10

20

30

0 20 40 60

Age (years)

Dep

th (

cm)

125

Appendix D (continued)

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

E

0

10

20

30

40

50

0.00 50.00 100.00 150.00


Dep

th (

cm)


F

0

10

20

30

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

Dep

th (

cm)


G

0

10

20

30

40

50

1.00 10.00 100.00


Dep

th (

cm)

H

0

10

20

0 20 40 60 80 100

Age (years)

Dep

th (

cm)

I

0

10

20

30

1.00 10.00 100.00

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

Dep

th (

cm)

126

Appendix D

Fig D-3.

J

0

10

20

30

40

50

0.00 20.00 40.00 60.00 80.00


Dep

th (

cm)


K

0

10

20

30

40

50

0.00 20.00 40.00 60.00 80.00

Ra-226 Activity (Bq/kg)

Dep

th (

cm)


L

0

10

20

30

40

50

1.00 10.00 100.00


Dep

th (

cm)

M

0

10

20

30

0 20 40 60 80

Age (years)

Dep

th (

cm)

Core 1 profile of A) Po210

B) Ra210

C) excess Pb210

activity and D) age

versus depth

Documents

Distribution and Speciation of Heavy Metals in Sediments from … · 2019-06-24 · Sharron, Sue, Burhan, Maree and Gavin for the moral support and encouragement. My thanks are due