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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
2
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)
3
Table of Contents
ABBREVIATIONS..................................................................9
ABSTRACT............................................................................10
CHAPTER I. INTRODUCTION ........................................12
1.1. BACKGROUND ..............................................................12
1.2. LAKE BURRAGORANG AND ITS CATCHMENT ...............14
1.3. REPORT ORGANISATION...............................................26
CHAPTER II. MATERIALS AND METHODS................28
2.1 FIELD SAMPLING .............................................................28
2.2 SEDIMENT GRAB .............................................................28
2.3 SEDIMENT CORE..............................................................31
2.4 ANALYTICAL METHODS..................................................33
2.4.1 MOISTURE CONTENT....................................................33
2.4.2 ORGANIC MATTER AND CARBONATE CONTENT ...........34
2.4.3 TOTAL NITROGEN AND PHOSPHORUS ..........................34
2.4.4 ACID EXTRACTABLE METAL........................................35
2.4.5 SPECIATION ..................................................................35
2.4.5.1 SEQUENTIAL EXTRACTION ........................................35
2.4.5.2 SIMULTANEOUSLY EXTRACTED METAL (SEM) AND
ACID VOLATILE SULPHIDE (AVS) ........................................36
2.4.6 SEDIMENTATION STUDY ..............................................39
2.4.7 STATISTICAL TREATMENT OF DATA ............................40
CHAPTER III. DISTRIBUTION OF METALS AND
SPECIATION IN SEDIMENT OF LAKE
BURRAGORANG USING SEQUENTIAL EXTRACTION
.................................................................................................42
3.1 INTRODUCTION................................................................42
3.2 STUDY AREA ...................................................................48
3.3 RESULTS AND DISCUSSION..............................................48
3.3.1 METAL DISTRIBUTION .................................................48
3.3.2 METAL SPECIATION .....................................................52
CHAPTER IV. DISTRIBUTION OF HEAVY METALS
AND THEIR BIOAVAILABILITY USING SEM AND AVS
IN THE SEDIMENTS OF LAKE BURRAGORANG .......61
5
4.1 INTRODUCTION................................................................61
4.2 STUDY AREA ...................................................................63
4.3 RESULTS AND DISCUSSION..............................................64
4.3.1 ORGANIC MATTER AND CARBONATE CONTENT ...........64
4.3.2 NUTRIENTS...................................................................65
4.3.3 BACKGROUND AND METAL DATA ...............................66
4.3.4 METALS........................................................................69
4.3.5 ACID VOLATILE SULPHIDE AND SIMULTANEOUSLY
EXTRACTED METALS ............................................................74
CHAPTER V. SEDIMENTARY RECORD OF HEAVY
METAL POLLUTION OF LAKE BURRAGORANG
USING 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 RESULTS AND DISCUSSION..............................................82
5.6.1 CORE 1 (NEAR DAMWALL) ...........................................82
5.6.2 CORE 2 (NEAR COX RIVER) ..........................................83
5.6.3 CORE 3 (NEAR NATTAI RIVER) .....................................83
CHAPTER VI. CONCLUSION ..........................................90
REFERENCES ......................................................................95
APPENDIX A.......................................................................114
APPENDIX B .......................................................................115
APPENDIX C.......................................................................119
APPENDIX D.......................................................................124
6
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
Table 5.2. Activity variation of 210
Po, 226
Ra and excess 210
Pb with depth in
sediment core 2 ................................................................................................. 86
Table 5.3. Activity variation of 210
Po, 226
Ra and excess 210
Pb with depth in
sediment core 3 ................................................................................................. 86
Table A-1 Uncertainty measurements for different studied variables .......... 114
7
List of Figures
Fig-1.1. Warragamba catchment showing Lake Burragorang [SCA, 1999] ..... 16
Fig-2.1. Locations of sediment core and grab samples in Lake Burragorang... 29
Fig 2.2. Ponar Petite sediment grab sampler........................................................ 30
Fig 2.3. Sediment grab sample collected from Lake Burragorang..................... 30
Fig 2.4. A) KB Sediment corer B) Sediment in an acrylic sediment core tube31
Fig 2.5. Sediment core extrusion device ................................................................ 32
Fig 2.6. Top of sediment core stripper .................................................................. 32
Fig 2.7. Details of sediment core stripper ............................................................. 33
Fig 2.8. Flow chart of sequential extraction scheme for sediments metal
speciation .......................................................................................................... 37
Fig 2.9. Extruding a sediment core in a glove box under nitrogen..................... 38
Fig. 3.1. The concentration of metals in the sediment grabs from Lake
Burragorang ..................................................................................................... 51
Fig 3.2. Metal distributions in Lake Burragorang sediments determined by
sequential extractions ...................................................................................... 55
Fig 3.3. Metal distributions in Lake Burragorang sediments determined by
sequential extractions ...................................................................................... 56
Fig 3.4. Metal distributions in Lake Burragorang sediments determined by
sequential extractions ...................................................................................... 57
Fig 3.5. Metal distributions in Lake Burragorang sediments determined by
sequential extractions ...................................................................................... 58
Fig 4.1. AVS and SEM distribution with depth ................................................... 77
Fig 5.1. Pathways by which 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
Fig 5.3. Lake Burragorang Core 2 age versus 1) Rainfall 2) Metals 3) Organic
matter and Carbonate content 4) Nutrients, Fe and Mn ............................. 88
Fig 5.4. Lake Burragorang Core 3 age versus 1) Rainfall 2) Metals 3) Organic
matter and carbonate content 4) Nutrients, Fe and Mn............................... 89
8
Fig B-1. Depth distributions of carbonate content, organic matter and
nutrients in sediments.................................................................................... 115
Fig B-2. Depth distributions of carbonate content, organic matter and
nutrients in sediments.................................................................................... 116
Fig B-3. Depth distributions of carbonate content, organic matter and
nutrients in sediments.................................................................................... 117
Fig B-4. Depth distributions of carbonate content, organic matter and
nutrients in sediments.................................................................................... 118
Fig C-1. Depth profiles of metals in sediments................................................. 119
Fig C-2. Depth profiles of metals in sediments................................................. 120
Fig C-3. Depth profiles of metals in sediments ................................................ 121
FigC-4. Depth profiles of metals in sediments.................................................. 122
Fig C-5. Depth profiles of metals in sediments................................................. 123
Fig D-1. Core 1 profile of A) Po210
B) Ra210
C) excess Pb210
activity and D) age
versus depth.................................................................................................... 124
Fig D-2. Core 2 profile of E) Po210
F) Ra210
G) excess Pb210
activity and H) age
I) excess Pb210
activity normalised with <63 μm size versus depth ........... 125
Fig D-3. Core 1 profile of A) Po210
B) Ra210
C) excess Pb210
activity and D) age
versus depth.................................................................................................... 126
9
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
10
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
11
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.
12
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
14
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
15
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
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
sites
0800
1600240032004000
30000
40000
50000
Con
cent
ratio
n (m
g/K
g)
Fe Mn
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
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
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Per
cent
age
(%)
Fra
ctio
n of
As
0
20
40
60
80
100
120
55
Fig 3.2. Metal distributions in Lake Burragorang sediments determined by
sequential extractions
Exchangeble Carbonate Fe-Mn Oxide Organic Inert
Sites
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Per
cent
age
(%)
Fra
ctio
n of
Cd
0
20
40
60
80
100
120
Sites
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
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
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Sites
Per
cent
age
(%)
Fra
ctio
n of
Cu
0
20
40
60
80
100
120
Fig 3.3. Metal distributions in Lake Burragorang sediments determined by
s sequential extraction
Exchangeble Carbonate Fe-Mn Oxide Organic Inert
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Sites
Per
cent
age
(%)
Fra
ctio
n of
Fe
0
20
40
60
80
100
120
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Sites
Sites
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Per
cent
age
(%)
Fra
ctio
n of
Mo
0
20
40
60
80
100
120
Exchangeble Carbonate Fe-Mn Oxide Organic Inert
Sites
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Per
cent
age
(%)
Fra
ctio
n of
Ni
0
20
40
60
80
100
120
Sites
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Per
cent
age
(%)
Fra
ctio
n of
Mn
0
20
40
60
80
100
120
57
Fig 3.4. Metal distributions in Lake Burragorang sediments determined by
sequential extractions
.
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
sequential extractions
Sites
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Per
cent
age
(%)
Fra
ctio
n of
V
0
20
40
60
80
100
120
Exchangeble Carbonate Fe-Mn Oxide Organic Inert
Sites
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
Per
cent
age
(%)
Fra
ctio
n of
Zn
DWA3 DWA2 DWA9 DWA12 DWA18 DWA19 DWA27 DWA35 DWA39 M3 M7
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 Results and Discussion
4.3.1 Organic Matter and carbonate content
The percentage of organic matter and carbonate content is given in Table 4.2 and
depth profiles of changes are shown in Figs B1-B4. In general carbonate contents
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
SEM Profile at DWA27
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 Profile at DWA27
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
SEM Profile at DWA35
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
AVS Profile at DWA35
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 Results and Discussion
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 Nin˜o episodes while periods
La Nin˜a episodes. Most El Nin˜o 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 Nin˜a events are
associated with above-average rainfall and flooding [Power, 2000]. The most
important episodes of La Niña or high rainfall occurred between 1945-1956, 1961 -
1969, 1974, 1976, 1978, 1984, 1988-1990 and a moderate La Niña event occurred in
1998/99, which weakened back to neutral conditions before reforming for a shorter
period in 1999/2000 causing widespread flooding throughout Australia [Bureau of
Meteorology, 2006].
Correlation was made unto 25 cm depth in core 1
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.
Table 5.1. Activity variation of 210
Po, 226
Ra and excess 210
Pb with depth in
sediment core 1
86
Table 5.2. Activity variation of 210
Po, 226
Ra and excess 210
Pb with depth in
sediment core 2
Table 5.3. Activity variation of 210
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
Fig 5.2. Lake Burragorang Core 1 age versus 1) Rainfall 2) Metals 3) Organic
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
Concentration (mg/kg)
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
% Organic Matter % Carbonate % moisture
0
5
10
15
20
25
30
35
0 10 20 30 40 50 60 70 80 90
Concentration (mg/kg)
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
Concentration (mg/kg)
Dep
th (
cm)
1935
1945
1955
1965
1975
1985
1995
2005
0 500 1000 1500 2000
Rainfall (mm)
Yea
r
Rainfall Bushfire
88
Fig 5.3. Lake Burragorang Core 2 age versus 1) Rainfall 2) Metals 3) Organic
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
% Organic Matter % Carbonate % moisture
0
5
10
15
20
25
30
35
0 10 20 30 40 50 60 70 80
Concentration (mg/kg)
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
Concentration (mg/kg)
Dep
th (
cm)
1935
1945
1955
1965
1975
1985
1995
2005
0 500 1000 1500 2000
Rainfall (mm)
Yea
r
Rainfall Bushfire
89
Fig 5.4. Lake Burragorang Core 3 age versus 1) Rainfall 2) Metals 3) Organic
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
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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
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
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
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
116
Appendix B (continued)
UWS15DWA27
Variables Variables
Fig B-2. Depth distributions of carbonate content, organic matter and
nutrients in sediments
0 5 10 15 20 500 1000 1500 2000
Dep
th (
cm)
0
10
20
30
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
0 5 10 15 300 600 900 1200 1500
Dep
th (
cm)
0
5
10
15
20
25
30
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
DWA30 UWS14Variables Variables
0 5 10 15 350 700 1050 1400 1750 2100
Dep
th (
cm)
0
10
20
30
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
0 7 14 21 28 35 600 1200 1800 2400 3000 3600
Dep
th (
cm)
0
5
10
15
20
25
30
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
117
Appendix B (continued)
Fig B-3.
DWA39DWA35
VariablesVariables
Depth distributions of carbonate content, organic matter and
nutrients in sediments
0 5 10 15 20350 700 1050 1400 1750
Dep
th (
cm)
0
5
10
15
20
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
0 5 10 15 400 800 1200 1600 2000 2400
Dep
th (
cm)
0
10
20
30
40
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
UWS13 DWA12Variables Variables
0 5 10 15 200 400 600 800 1000120014001600
Dep
th (
cm)
0
10
20
30
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
0 5 10 15 20 500 1000 1500 2000 2500 3000 3500
Dep
th (
cm)
0
10
20
30
40
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
118
Appendix B
Fig B-4.
DWA18 MC3Variables Variables
Depth distributions of carbonate content, organic matter and
nutrients in sediments
0 5 10 15 20 500 1000 1500 2000 2500
Dep
th (
cm)
0
5
10
15
20
25
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
0 5 10 15 300 600 900 1200 1500 1800 2100 2400
Dep
th (
cm)
0
10
20
30
40
TP (mg/Kg)TN (mg/Kg)% Organic matter% Carbonate
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
Metal (mg/Kg) Metal (mg/Kg)
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.
Depth profiles of metals in sediments
Metal (mg/Kg)Metal (mg/Kg)
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
Metal (mg/Kg)Metal (mg/Kg)
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
Metal (mg/Kg)Metal (mg/Kg)
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
Appendix C (continued)
Fig C-3.
121
Depth profiles of metals in sediments
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
Metal (mg/Kg) Metal (mg/Kg)
DWA30
350400450500550 28000 32000 36000
Dep
th (
cm)
0
10
20
30
0 20 40D
epth
(cm
)60 80
0
10
20
30
40
Metal (mg/Kg)Metal (mg/Kg)
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
Metal (mg/Kg) Metal (mg/Kg)
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
Appendix C (continued)
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
Metal (mg/Kg) Metal (mg/Kg)
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
Metal (mg/Kg) Metal (mg/Kg)
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
Depth profiles of metals in sediments
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)
Vertical line = average Ra-226
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
Po-210 Activity (Bq/kg)
Dep
th (
cm)
Vertical line = average Ra-226
F
0
10
20
30
0.00 50.00 100.00 150.00Ra-226 Activity (Bq/kg)
Dep
th (
cm)
Vertical line = average Ra-226
G
0
10
20
30
40
50
1.00 10.00 100.00
Pb-210 (excess) Activity (Bq/kg)
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
Po-210 Activity (Bq/kg)
Dep
th (
cm)
Vertical line = average Ra-226
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)
Vertical line = average Ra-226
L
0
10
20
30
40
50
1.00 10.00 100.00
Pb-210 (excess) Activity (Bq/kg)
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