Bethany Megan O'Shea - UNSWorks - UNSW Sydney

DELINEATING THE SOURCE,

GEOCHEMICAL SINKS AND AQUEOUS

MOBILISATION PROCESSES OF

NATURALLY OCCURRING ARSENIC IN A

COASTAL SANDY AQUIFER

Stuarts Point, New South Wales, Australia.

Bethany Megan O’Shea

A thesis submitted in fulfillmentof the requirements for the degree of

Doctor of Philosophy

School of Biological, Earth & Environmental Sciences Faculty of Science

University of New South Wales Sydney, Australia

August 2006

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Surname or Family name: O'SHEA

THE UNIVERSITY OF NEW SOUTH WALES Thesis/Dissertation Sheet

First name: BETHANY Other name/s: MEGAN

Abbreviation for degree as given in the University calendar: PhD

School: BIOLOGICAL, EARTH & ENVIRONMENTAL SCIENCES Faculty: SCIENCE

Title: DELINEATING THE SOURCE, GEOCHEMICAL SINKS AND AQUEOUS MOBILISATION PROCESSES OF NATURALLY OCCURRING ARSENIC IN A COASTAL SANDY AQUIFER, STUARTS POINT, NEW SOUTH WALES, AUSTRALIA

Abstract 350 words maximum: (PLEASE TYPE)

Elevated arsenic concentrations have been reported in a drinking water and irrigation-supply aquifer of Stuarts Point, New South Wales, Australia. Arsenic occurrence in such aquifers is potentially a major issue due to their common use for high yield domestic and irrigation water supplies. Ten multi-level piezometers

were installed to depths of approximately 30 m in the sand and clay aquifer. Sediment samples were collected at specific depths during drilling and analysed for chemical and mineralogical composition, grain size characteristics, potential for arsenic release from solid phase and detailed microscopic features. From this data, a full geomorphic reconstruction allowed the determination of source provenance for the aquifer sediments. The model proposed herein provides evidence that the bulk of the aquifer was deposited under

intermittent fluvial and estuarine conditions; and that all sediments derive from the regional arsenicmineralised hinterland. More than 200 groundwater samples were collected and analysed for over 50

variables. The heterogeneity of the aquifer sediments causes redox stratification to occur, which in turn governs arsenic mobility in the groundwater. The bulk of the aquifer is composed of fluvial sand deposits

undergoing reductive dissolution of iron oxides. Arsenic adsorbed to iron oxide minerals is released during dissolution but re-adsorbs to other iron oxides present in this part of the aquifer. The deeper, more

reducing fluvial sand and estuarine clay groundwaters have undergone complete reductive dissolution of iron oxides resulting in the subsequent mobilisation of arsenic into groundwater. Some of this arsenic has

been incorporated into iron sulfide mineral precipitates, forming current arsenian pyrite sinks within the aquifer. The extraction of groundwater from the aquifer for irrigation and drinking water supply induces seawater intrusion of arsenic-rich estuarine water, bringing further dissolved arsenic into the aquifer. A greater understanding of the source, sinks and mobilisation of arsenic in this aquifer contributes to our

broad understanding of arsenic in the environment; and allows aquifer specific management procedures and research recommendations to be made. Any coastal or unconsolidated aquifer that has sediments

derived from mineralised provenances should consider monitoring for arsenic, and other potentially toxic trace elements, in their groundwater systems.

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O’Shea (2006) Page i

ABSTRACT

Elevated arsenic concentrations have been reported in a drinking water and irrigation-

supply aquifer of Stuarts Point, New South Wales, Australia. Coastal areas are often

heavily populated, making groundwater a valuable resource. Arsenic occurrence in such

aquifers is potentially a major issue due to their common use for high yield domestic

and irrigation water supplies. This study was initiated in response to community anxiety

over arsenic in their water supply aquifer; and concern for the potential occurrence of

elevated arsenic in other coastal sandy aquifers.

The source of arsenic deduced from preliminary studies was proposed as being derived

from acid sulfate soil material and sorption of arsenic onto marine clays during sea level

transgressions in the Quaternary period. The evidence presented herein suggests an

alternative source; the arsenic in this aquifer is predominantly derived from regional

and/or bedrock geology.

Ten multi-level piezometers were installed to depths of approximately 30 m in the sand

and clay aquifer. Sediment samples were collected at specific depths during drilling and

analysed for chemical and mineralogical composition, grain size characteristics,

potential for arsenic release from solid phase and detailed microscopic features. From

this data, a full geomorphic reconstruction allowed the determination of source

provenance for the aquifer sediments. Previous geomorphic models have suggested the

aquifer is largely comprised of barrier sands derived from off-shore. The model

proposed herein provides evidence that the bulk of the aquifer was deposited under

intermittent fluvial and estuarine conditions; and that all sediments derive from the

regional arsenic-mineralised hinterland. The wet climate of the Pleistocene and

Holocene promoted weathering in the upper catchment, which weathered natural

stibnite deposits containing arsenic and antimony to form iron oxides. These oxides

were present in colloidal forms and as coatings on sediment grains, which were

transported downstream and deposited to form the current aquifer matrix.

Abstract

O’Shea (2006) Page ii

More than 200 groundwater samples were collected and analysed for over 50 variables.

Statistical methods were employed in conjunction with hydrochemical interpretative

techniques to determine the geochemical processes active in the groundwater system.

Natural and anthropogenic effects contribute to arsenic mobilisation. The heterogeneity

of the aquifer sediments causes redox stratification to occur, which in turn governs

arsenic mobility in the groundwater. Shallow groundwaters are exposed to nitrate input

from the ground surface, which contributes to the oxidative release of discrete arsenian

pyrite phases. The bulk of the aquifer is composed of fluvial sand deposits undergoing

reductive dissolution of iron oxides. Arsenic adsorbed to iron oxide minerals is released

during dissolution but re-adsorbs to other iron oxides present in this part of the aquifer.

The deeper, more reducing fluvial sand and estuarine clay groundwaters have

undergone complete reductive dissolution of iron oxides resulting in the subsequent

mobilisation of arsenic into groundwater. Some of this arsenic has been incorporated

into iron sulfide mineral precipitates, forming current arsenian pyrite sinks within the

aquifer. Mineralised bedrock groundwaters contribute arsenic to the overlying

weathered bedrock clays. The extraction of groundwater from the aquifer for irrigation

and drinking water supply induces seawater intrusion of arsenic-rich estuarine water,

bringing further dissolved arsenic into the aquifer.

Previous arsenic mobilisation processes were proposed to be dominated by the

dissolution of aluminium hydroxides and release of adsorbed As(V) or concurrently via

pH-influenced desorption of arsenic enriched iron oxides; and the leaching of the

aquifer’s sandy-clayey matrix by groundwaters rich in HCO3- and of high pH. The

detailed hydrochemical interpretation provided herein suggests these are minor arsenic

control mechanisms. Instead, the reductive dissolution of iron oxides and precipitation

of iron sulfide minerals has been shown to dominate arsenic mobilisation and

retardation at Stuarts Point.

A greater understanding of the source, sinks and mobilisation of arsenic in this aquifer

contributes to our broad understanding of arsenic in the environment; and allows aquifer

specific management procedures and research recommendations to be made. In

addition, the findings of this study are similar to geochemical processes proposed by

some authors for arsenic release in Bangladesh. The suggestion that groundwater

Abstract

O’Shea (2006) Page iii

arsenic occurrence has implications for the management of coastal aquifers remains, but

not solely due to the presence of acid sulfate soils or exposure of these coastal

sediments to Quaternary sea level fluctuations. Rather, any coastal or unconsolidated

aquifer that has sediments derived from mineralised provenances should consider

monitoring for arsenic, and other potentially toxic trace elements, in their groundwater

systems.

O’Shea (2006) Page iv

ACKNOWLEDGEMENTS

It is good to have an end to journey towards, but it is the journey that matters, in the end.

– Ursula Le Guin

There are many who have supported me during my journey towards a PhD. Some have

been there from the first step, while others have endured the final leg of the journey.

Each of you have helped in your own special way, and for this, I thank you.

I would like to acknowledge my supervisor Jerzy Jankowski who inspired my interest in

hydrogeochemistry and Jesmond Sammut who provided both advice and

encouragement. The day to day trials and successes were experienced by those present

in room 515 of the School of BEES – Sarah Groves, John Wischusen, Jessica Northey,

Maria Dubikova, Emma Haradasa and Karina Morgan. Our many conversations – both

hydrochemical and ‘tangent-related’ – were excellent stress relievers (in addition to

coffee and chocolate). Our combined quest for knowledge saw us endlessly searching

through journal papers and text books to find ‘the right answer’ even when there may

not have been one. I have learnt a great deal from these brainstorming sessions.

To those that helped me in the field or lab – Kavita Gosavi, Dorothy Yu, Irene

Wainwright, Lange Jorstad, James Smith, Phillip Crisp, Maree Emmett, Paul Simpson,

Barry Searle – your help is greatly appreciated. And to those who stood on the outside

looking in – my family and friends. Thankyou for your cooking (Mum and Alana),

outrageous yet funny technical suggestions (Dad), constant asking of when my ‘book’

will be finished (everyone single person I know!!), nights on the town (Melbourne girls,

Coogee boys) and the combined support, nagging and encouragement from Jim.

To all those who have contributed to the development of my scientific curiosity, I hope

this thesis inspires you all. But most of all…..

Happy Reading !!

O’Shea (2006) Page v

TABLE OF CONTENTS

Abstract i Acknowledgements iv Table of Contents v List of Figures xi List of Tables xv Abbreviations xvii

1 INTRODUCTION 1 1.1 PROJECT INCEPTION 1

1.2 RESEARCH AIMS 3

1.3 HYPOTHESES 3

1.4 SIGNIFICANCE OF THIS RESEARCH 4

1.5 STUDY APPROACH 6

1.6 THESIS STRUCTURE 6

2 THE STUARTS POINT STUDY AREA 9 2.1 LOCATION AND ENVIRONMENTAL CONDITIONS 9

2.2 THE MACLEAY RIVER CATCHMENT 9

2.2.1 Regional Geology 12

2.2.2 The Macleay Fluvial-Deltaic Floodplain 15

2.3 DEPOSITION OF THE STUARTS POINT AQUIFER 15

2.3.1 Previous Geomorphic Models 15

2.3.2 Bedrock Geology 19

2.3.2.1 Lithology and Structure 19

2.3.2.2 Mineralisation 22

2.3.3 Hydrogeology 25

2.4 PREVIOUS INVESTIGATIONS 28

3 ARSENIC LITERATURE REVIEW 31 3.1 GENERAL PROPERTIES OF ARSENIC 31

3.1.1 Periodicity 31

3.1.2 Toxicity 33

3.1.3 Arsenic Production and Use 34

3.2 SOURCES OF ARSENIC IN THE ENVIRONMENT 35

3.2.1 Natural Occurrences 36

3.2.2 Anthropogenic Sources 37

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O’Shea (2006) Page vi

3.2.3 Concentrations in the Environment 39

3.2.3.1 Rocks 42

3.2.3.2 Soils 42

3.2.3.3 Sediments 42

3.2.3.4 Water 42

3.2.3.5 Air 43

3.3 AQUEOUS ARSENIC CHEMISTRY 44

3.4 IMMOBILISATION OF ARSENIC: SOLID PHASE RETARDATION 47

3.4.1 Properties of Sorbent Materials 47

3.4.1.1 Surface Area 47

3.4.1.2 Surface Charge 47

3.4.2 Surface Sorption 49

3.4.2.1 Surface Complexation 49

3.4.2.2 Permanent Versus Variable Surface Charge 51

3.4.3 Arsenic Retention in the Solid Phase 52

3.4.3.1 Solid Phase Arsenic Formation 52

3.4.3.2 Surface Reactions Affecting Arsenic Mobility 56

3.4.3.2.1Surface Complexation (Adsorption) 56

3.4.3.2.2Surface Precipitates 65

3.4.3.2.3Competitive Anion Exchange 65

3.5 MOBILISATION OF ARSENIC: THE GEOCHEMICAL TRIGGERS 67

3.5.1 Changes to Aqueous pH 67

3.5.2 Shifts in Aqueous Redox Potential 68

3.5.3 Influence of, and Interaction with, the Surrounding Solution 70

3.5.4 Colloidal Transport 72

3.6 NATURALLY OCCURRING GROUNDWATER ARSENIC 72

3.6.1 Geochemical Controls 74

3.6.1.1 Oxidising Environments 74

3.6.1.2 Arid Oxidising Environments 75

3.6.1.3 Reducing Conditions 76

3.6.1.4 Combined Oxidising and Reducing Conditions 77

3.6.2 Geological And Depositional Influences 78

3.6.2.1 Geothermal Systems 78

3.6.2.2 Volcanic Sediments 78

3.6.2.3 Alluvial and Deltaic Sediments 79

3.6.2.4 Consolidated Sediments 79

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O’Shea (2006) Page vii

3.6.2.5 Lacustrine Environments 80

3.6.2.6 Glacial Drift 81

3.6.2.7 Aeolian Loess Deposits 82

3.6.2.8 Zone of Water Table Fluctuation 82

3.6.2.9 Coastal Sand Dunes 83

3.6.3 Arsenic In Australia 84

3.7 ARSENIC GEOCHEMICAL SUMMARY 84

4 METHODOLOGY 86 4.1 CONSTRUCTION & INSTALLATION OF SAMPLING NETWORK 86

4.1.1 Multi-Level Piezometer Design 86

4.1.2 Drilling 86

4.2 SEDIMENT AND GROUNDWATER SAMPLING 88

4.2.1 Sediment Sampling, Storage and Preparation Methods 88

4.2.2 Groundwater Sampling, Preservation & Storage Methods 88

4.3 AQUEOUS CHEMICAL ANALYSES 89

4.3.1 General Parameters 89

4.3.2 Unstable Chemical Species 90

4.3.3 Major Ions 90

4.3.4 Minor Elements 91

4.3.5 Arsenic Speciation 92

4.4 SOLID PHASE ANALYSES 93

4.4.1 Electron Microscopy 93

4.4.1.1 Sample Preparation 93

4.4.1.2 SEM-EDS 93

4.4.1.3 Electron Microprobe-WDS 94

4.4.1.4 Internal QA/QC Procedures 94

4.4.2 Grain Size Analysis 95

4.4.3 Sequential Extractions 97

4.4.4 XRD 97

4.4.5 XRF 97

4.4.6 Loss On Ignition (LOI) 100

4.5 STATISTICAL ANALYSES 100

4.5.1 Data Screening 100

4.5.1.1 Assessment of Normality 100

4.5.1.2 Standardization 101

Table of Contents

O’Shea (2006) Page viii

4.5.1.3 Censored Data 102

4.5.2 Descriptive Statistics 103

4.5.3 Multi-Variate Statistical Analyses 104

4.5.3.1 Cluster Analysis 104

4.5.3.2 Principal Components Analysis (PCA) 104

4.5.4 Statistical Correlation 105

5 ORIGINAL SOURCE OF ARSENIC TO THE AQUIFER 106 5.1 HYPOTHESES 106

5.2 INVESTIGATIVE RESULTS AND INTERPRETATION 109

5.2.1 Aquifer Lithology 109

5.2.2 Aquifer Depositional Conditions 120

5.2.2.1 Differentiating between Beach Sand & Fluvial Sand 120

5.2.2.2 The Onset of Estuarine Conditions 122

5.2.2.3 Weathered Bedrock 126

5.2.3 Aquifer (Sediment) Chemistry 128

5.2.4 Aquifer Facies Definition 128

5.2.4.1 Statistical Procedures 128

5.2.4.2 Development of Aquifer Cross Sections 132

5.2.5 Determining Source Provenance for the Aquifer Facies 137

5.2.5.1 Origin of the Sands 137

5.2.5.2 Selection of a Hinterland Indicator Element 139

5.2.6 Aquifer Geomorphology 142

5.2.6.1 Linking the Stuarts Point Facies to Surrounding

Depositional Environments 142

5.2.6.2 Proposed Geomorphic Model for the Stuarts Point

Aquifer 143

5.3 DISCUSSION ON EACH PROPOSED ARSENIC SOURCE

HYPOTHESIS 146

5.3.1 Anthropogenic Sources 146

5.3.1.1 Agricultural Practices 147

5.3.1.2 Historical Mining 148

5.3.1.3 Town (Anthropogenic) By-Products 148

5.3.2 A Regional Geologic Source? 150

5.3.3 Input from As-rich Holocene Sea Level Rise 152

5.3.4 A Note on Arsenic and Acid Sulfate Soils / Pyrite 157

Table of Contents

O’Shea (2006) Page ix

5.3.5 Direct Bedrock Contribution 158

5.4 A COMPARISON TO ARSENIC SOURCES IN OTHER NATURALLY

ELEVATED ARSENIC ENVIRONMENTS 160

5.5 CHAPTER SUMMARY 162

6 IDENTIFICATION OF CURRENT ARSENIC SINKS 164 6.1 AQUIFER MINERALOGY AND PHYSICAL PROPERTIES 165

6.2 HOW MUCH ARSENIC IS IN THE AQUIFER MATRIX ? 167

6.2.1 Arsenic-Bearing Minerals 174

6.2.2 Are there any Distinctive Correlations? 175

6.3 EXAMINATION OF KNOWN ARSENIC SCAVENGING MATERIALS IN

RELATION TO THE STUARTS POINT AQUIFER MATRIX 181

6.3.1 Adsorption to Oxyhydroxides 181

6.3.2 Association with Pyrite 188

6.3.3 Incorporation into Clay Minerals 194

6.3.4 Sorption on Organic Matter 196

6.3.5 Association with Calcite 197

6.3.6 Anion Competition 198

6.4 CHAPTER SUMMARY: IDENTIFICATION OF CONTROLLING SINKS

IN THE AQUIFER 198

7 ARSENIC MOBILISATION IN THE AQUIFER 201 7.1 GENERAL AQUIFER CONDITIONS 201

7.1.1 Major Hydrochemical Processes 201

7.2 ARSENIC GEOCHEMICAL PROCESSES 207

7.2.1 Arsenic Distribution and Speciation 207

7.2.2 Identification of Arsenic Mobilisation Processes 211

7.2.2.1 Shallow Groundwaters 212

7.2.2.2 Barrier Sand Groundwaters 220

7.2.2.3 Fluvial Sand Groundwaters 223

7.2.2.4 Fluvial Sand / Estuarine Clay Groundwaters 231

7.2.2.5 Bedrock Clay Groundwaters 236

7.2.2.6 Seawater Intrusion Groundwaters 237

7.3 CHAPTER SUMMARY 241

Table of Contents

O’Shea (2006) Page x

8 CONCLUSIONS AND RECOMMENDATIONS 243 8.1 THE PROBLEM 243

8.2 PROPOSED ARSENIC GEOCHEMICAL MODEL FOR THE STUARTS

POINT AQUIFER 243

8.3 CONCLUSIONS 246

8.4 RECOMMENDATIONS 248

8.4.1 Aquifer Specific (Stuarts Point) 248

8.4.2 Other Research Recommendations 251

REFERENCES 254

APPENDICES A PUBLICATIONS

B1 SEDIMENT CHEMISTRY

B2 GROUNDWATER CHEMISTRY

B3 GRAIN SIZE ANALYSIS

B4 BORE LOGS

B5 IRON OXIDE CONVERSIONS

B6 SEQUENTIAL EXTRACTION RESULTS

B7 STATISTICAL RESULTS

B8 GEOCHEMICAL MODELLING RESULTS

C1 PHOTOGRAPHS

O’Shea (2006) Page xi

LIST OF FIGURES

Figure 1.1 Thesis Structure………………………………………………….…….....8Figure 2.1 Location of the Stuarts Point sand aquifer (contained within the dotted

lines). Multi-level piezometers used in this study are also shown…..…..10Figure 2.2 Boundaries of the Macleay River Catchment. Stuarts Point is located on

the northeastern margin of the catchment where floodplain deposits dominate (NSW EPA, 2005)………………………………………….....11

Figure 2.3 Extent of the New England Fold Belt and its adjacent basin margins (modified from Palfreyman, 1984 and Leitch, 1974)……………….…..12

Figure 2.4 Geological subdivision of the southern NEFB. Stuarts Point is situated on the Nambucca Block (modified from Gilligan et al., 1992)……………13

Figure 2.5 Generalised section of a dual barrier system (Roy and Thom, 1981)…..16Figure 2.6 Pre-Pleistocene and current Holocene coastlines on the NSW mid-north

coast (Hails, 1968)……………………………………………………....17Figure 2.7 The Stuarts Point aquifer denoted by Eddie (2000) as an inner barrier

beach/dune sand – ‘sp’ (extracted and modified from Eddie, 2000)……18Figure 2.8 Various bedrock lithologies proposed for Stuarts Point. The Stuarts Point

aquifer is contained within the dotted area……………………………...20Figure 2.9 Coastal granitoid belt forming Mt Yarrahapinni (modified from Gilligan

et al., 1992)………………………………………………………..…….23Figure 2.10 Zones of mineralisation surrounding Mt Yarrahapinni (modified from

Gilligan et al., 1992). Numbers refer to recognised mineral deposits described in Gilligan et al. (1992). 772-774 are disseminated/vein Mo deposits……………………………………………………………….…24

Figure 2.11 Groundwater contour map (mAHD) for the Stuarts Point sand aquifer…………………………………………………………………..26

Figure 2.12 Regional arsenic distribution reported by Smith et al.(2000).The Stuarts Point aquifer and multi-level piezometers installed for this study are marked…………………………………………………………………..29

Figure 3.1 Eh-pH diagram for the As-O2-S-H2O system at 25ºC and 1 bar pressure (Bissen and Frimmel, 2003)…………………………………..…………45

Figure 3.2 Deprotonation of (a) As(III) and (b) As(V) species over different pH ranges (Smedley and Kinniburgh, 2002)………………………………..46

Figure 3.3 Mechanisms of sorption at the mineral/water interface. (a) adsorption of an ion via formation of an outer-sphere complex; (b) loss of hydration water and formation of an inner-sphere complex; (c) lattice diffusion and isomorphic substitution within the mineral lattice; (d) rapid lateral diffusion and formation either of a surface polymer or (e) adsorption on a ledge which maximizes the number of bonds to the atom; (f) upon particle growth surface polymers end up embedded in the lattice structure; and (g) the adsorbed ion can diffuse back into solution either as a result of dynamic equilibrium or as a product of surface reactions (from Charlet and Manceau, 1993; as cited in Sparks, 2003)…………………………..49

Figure 3.4 (a) Inner-sphere complex formation and (b) Outer-sphere complex formation (taken from Sposito, 1984; as cited in Sparks, 2003)…….….50

Figure 3.5 Arsenic adsorption on HFO for a) As(V) and b) As(III) according to ionic strength (I), Fe and As concentrations given (after Dzombak and Morel,

List of Figures

O’Shea (2006) Page xii

1990). Solid line is optimal fit, dashed line is an estimate using sorption constants……………………………………………………………….59

Figure 3.6 As(III)substitution for the carbonate ion on the calcite surface as suggested by Cheng et al. (1999)………………..…………………….64

Figure 3.7 Natural arsenic distribution in aquifers globally (Smedley and Kinniburgh, 2002). Some mining and geothermal sources are also noted………….73

Figure 4.1 Schematic of the multi-level design used at Stuarts Point……………..87Figure 5.1 Borelog for ML1………………………………………………………110Figure 5.2 Borelog for ML2………………………………………….…………...111Figure 5.3 Borelog for ML3………………………………………………………112Figure 5.4 Borelog for ML4………………………………………………………113Figure 5.5 Borelog for ML5………………………………………………………114Figure 5.6 Borelog for ML6………………………………………………………115Figure 5.7 Borelog for ML7………………………………………………………116Figure 5.8 Borelog for ML8………………………………………………………117Figure 5.9 Borelog for ML9………………………………………………………118Figure 5.10 Borelog for ML10…………………………………………………….119Figure 5.11 Moment skewness versus moment standard deviation to differentiate

beach and river sands from the Stuarts Point aquifer (method adapted from Friedman, 1960). Numbers represent depth below ground surface for respective multi-levels……………………………………………122

Figure 5.12 Proposed bedrock units at Stuarts Point. The Stuarts Point aquifer is contained within the dotted area. Approximate bedrock contours (based on drilling evidence) are shown at 5m depth intervals……………….127

Figure 5.13 The HCA dendrogram on sediment chemistry successfully delineated the proposed facies units…………………………………………………131

Figure 5.14 Cross Section A to A'. See text for discussion……………………….133Figure 5.15 Cross section B to B'. See text for discussion………………………..134Figure 5.16 Cross Section C to C'. See text for discussion……………………….135Figure 5.17 Cross section D to D'. See text for discussion………………………136Figure 5.18 Ternary diagram of the Stuarts Point sediments showing a recycled

orogen for most aquifer sediments. Boundaries for tectonic fields are from Dickinson and Suczek (1979), Dickinson et al. (1983), and Ingersoll and Suczek (1979) (Q=quartz, F=feldspar, L=lithics)………………138

Figure 5.19 Indicative types of mineralisation occurring within the Macleay River catchment (information obtained from Gilligan et al., 1992)……….140

Figure 5.20 Proposed geomorphic model for the deposition of the Stuarts Point aquifer. ……………………………………………………………….145

Figure 5.21 Potential anthropogenic arsenic sources in the Stuarts Point aquifer...146Figure 5.22 Sb versus As concentrations in the aquifer matrix. The “beach barrier

deposits” contain the most Sb and the least As, showing Sb is more immobile than As……………………………………………………..151

Figure 5.23 Vertical solid phase arsenic concentrations reported herein. Arsenic peaks with “fluvial sand/estuarine clay” units due to their higher adsorption capacity……………………………………………………………….156

Figure 5.24 As and Mo groundwater concentrations with depth in ML9 (left) and ML6 (right). Note the simultaneous increase in both elements as depth increases and weathered phyllite bedrock is penetrated at 20 mbgs in

List of Figures

O’Shea (2006) Page xiii

ML9. Contrasting, ML6 shows no distinct correlation between the two elements……………………………………………………………159

Figure 5.25 Development of an arsenic affected aquifer in Taiwan. Slate debris, deposited via fluvial and long-shore processes, is the source of the arsenic (Yu et al., 2000)……………………………………………161

Figure 6.1 SEM photomicrographs of typical grains observed for each geomorphic facies in the Stuarts Point aquifer. (a) beach barrier sand (b) fluvial sand (c) fluvial sand/estuarine clay and (d) bedrock clay. Spectral plots coincide with the white spectral boxes outlined in each photomicrograph……………………………………………………166

Figure 6.2 XRD spectrum for a fluvial sand/estuarine clay at ML7/27. Quartz dominates with minor amounts of illite, kaolinite, feldspar and pyrite identified……………………………………………………………174

Figure 6.3 As versus fine fraction component (silts and clays) in the aquifer sediments. The clayey units report the highest As concentrations yet no good correlation exists to support As association with just the fine fraction………………………………………………………………180

Figure 6.4 Detailed analysis of particles within oxidised bedrock clay (ML10/26.8) with some fluvial influence. Potential iron oxide minerals, as proposed by atomic ratios, are (a) goethite (b) ferrihydrite aggregate on goethite surface (c) amorphous iron oxide on goethite surface………………186

Figure 6.5 Small agregate observed on a “fluvial sand / estuarine clay” sample. Clay and quartz are suspected however the presence of sulfur may indicate discrete sulfide mineral presence……………………………………189

Figure 6.6 Electron microprobe back scattered image of a shell fragment with a pyrite cluster, as determined by the quantitative analysis…………...191

Figure 6.7 Back scattered electron image of a quartz grain with an arsenic containing pyrite cluster occurring on it…………………………………………193

Figure 6.8 A coating on a grain in sample ML10/26.8 under electron microprobe (a) optical image (b) back scattered electron image. Scale identical……194

Figure 6.9 Line scan output chart and identified mineral compositions…………195Figure 6.10 Coffee rock SEM photomicrograph………………………………….196Figure 6.11 Portion of shell material identified as calcite with minor trace element

impurities……………………………………………………………..197Figure 6.12 Summary of As geochemical sinks proposed herein for the heterogenous

Stuarts Point aquifer………………………………………………….200Figure 7.1 Piper diagram showing lithological and geochemical influences on major

ion groundwater composition. The log of ML1 is shown to illustrate how groundwater varies vertically down the profile………………………202

Figure 7.2 Saturation Indices for calcite with the various water groups separately graphed………………………………………………………………..203

Figure 7.3 Ca+Mg+Sr versus HCO3- to indicate calcite dissolution. The excess

cations over HCO3- in the seawater intrusion samples is a result of

mixing and cation exchange processes………………………………..204Figure 7.4 Na+ versus Cl- shows seawater intrusion and mixing groundwaters plot

on the seawater line (a). Other groundwaters (red box in {A} enlarged) are less influenced (b) by seawater composition……………………...205

List of Figures

O’Shea (2006) Page xiv

Figure 7.5 (a) Reverse ion exchange dominates the seawater intrusion samples whereas (b) the shallow groundwaters, barrier sands and fluvial sand and clay are subjected to both normal and reverse ion exchange processes.206

Figure 7.6 Stuarts Point groundwater samples plotted on a generalised Eh-pH diagram for the As-O2-H2O system at 25ºC and 1 bar total pressure....210

Figure 7.7 Vertical distribution of dissolved arsenic for ML1 to ML5. Statistically defined water groups are noted on each graph………………………..213

Figure 7.8 Vertical distribution of dissolved arsenic for ML6 to ML10. Statistically defined water groups are noted on each graph………………………..214

Figure 7.9 (a) Na+ versus Cl- (b) As Tot versus SO42- (c) Fe2+ versus SO4

2- and (d) AsTot versus Fe2+ concentrations in shallow groundwaters…………..215

Figure 7.10 (a) arsenic versus ammonium in the shallow groundwaters, and (b) iron versus ammonium in the shallow groundwaters………………………216

Figure 7.11 Plot of dissolved Al variation with depth for ML2. Statistical groundwater groups are noted also……………………………………218

Figure 7.12 Saturation Indices of various aluminium oxide minerals (boehmite, diaspore and gibbsite) for the shallow groundwaters………………….219

Figure 7.13 AsTot versus (a) Cl- (b) EC (c) TDS and (d) Li+ in the shallow groundwaters…………………………………………………………..220

Figure 7.14 (a) Iron versus arsenic (b) sulfate versus arsenic (c) arsenic versus ammonium and (d) iron versus ammonium in the barrier sand groundwaters…………………………………………………………..221

Figure 7.15 (a) Fe versus HCO3- (b) As versus HCO3

- (c) As versus Fe and (d) As versus Eh for the fluvial sand groundwaters…………………………..225

Figure 7.16 (a) As versus pH (b) PO43- versus As (c) As versus Mo (d) As versus Cr

(e) As versus V (f) As versus U for the fluvial sand groundwaters…...229Figure 7.17 As versus Sr in the fluvial sand groundwaters of Stuarts Point……….230Figure 7.18 Saturation indices for pyrite in the fluvial sand / estuarine clay

groundwaters. Ninety three percent of these waters are oversaturated.233Figure 7.19 Constituents required for pyrite formation (Appelo & Postma, 1999)..232Figure 7.20 AsTot, Na+, Cl- and EC variation with depth for ML1…………………237Figure 7.21 Arsenic versus chloride (a) arsenic versus sulfate (b) in the seawater

intrusion groundwaters………………………………………………...238Figure 7.22 The Macleay River exhibits increased arsenic concentrations

downgradient of the Hillgrove antimony mines (Ashley et al., 2003)...240Figure 8.1 Arsenic source and mobilisation model proposed for Stuarts Point…..245

O’Shea (2006) Page xv

LIST OF TABLES

Table 1.1 Tasks executed in this thesis and new data generated as a result.................... 7Table 2.1 Geological History of the NEFB. .................................................................. 14Table 3.1 Common Physico-Chemical Properties of Arsenic (adapted from Carmalt

and Norman, 1998; Moeller et al., 1989; Reimann and de Caritat, 1998). ... 32Table 3.2 Past and present uses of arsenical compounds (adapted from Nriagu &

Azcue, 1990; Bissen and Frimmel, 2003; WHO, 1992). .............................. 34Table 3.3 Common naturally occurring arsenic bearing minerals and their

occurrence in the environment (adapted from Smedley and Kinniburgh, 2002; Palache et al., 1949). ........................................................................... 36

Table 3.4 Reported arsenic concentrations in various media. ....................................... 40Table 3.5 Dissociation constants for protonated arsenite and arsenate. ........................ 44Table 3.6 Measured or estimated surface areas and surface site densities for

different geological materials (table adapted from Langmuir, 1997)............ 48Table 3.7 CEC’s of different geologic materials measured at pH 7. Absent,

negligible or slight pH dependencies indicate the material obtains its CEC primarily via isomorphous substitution. Strong dependencies indicate the surface charge results from adsorption at the mineral surface (table adapted from Langmuir, 1997)...................................................................... 51

Table 3.8 pH concentrations required for optimum arsenic sorption conditions onto different iron oxides. ..................................................................................... 58

Table 3.9 Surface complexation reactions for arsenate and arsenite on HFO (data derived from Dzombak and Morel, 1990)..................................................... 58

Table 3.10 Changes to redox equilibrium in natural waters and how this may affect arsenic mobilisation....................................................................................... 69

Table 3.11 Common colloids and their PZC (vanLoon and Duffy, 2000)...................... 72Table 4.1 Depth and screened intervals for the multi-levels used in this study. Note

that a screened interval of 5-30 indicates that a groundwater sample was collected at metre increments between 5 m depth and 30 m depth, unless otherwise noted.............................................................................................. 88

Table 4.2 Calculation of various grain size parameters using different methods.......... 96Table 4.3 Mean RPD values between calculations by the graphical and moment

measures methods.......................................................................................... 96Table 4.4 Step-by-step sequential extraction scheme utilised herein............................ 98Table 4.5 Commonly applied sequential extraction techniques and their associated

advantages/disadvantages.............................................................................. 99Table 5.1 Techniques used herein to dismiss or support potential arsenic sources to

the Stuarts Point aquifer, and the location of results provided within this thesis. ........................................................................................................... 107

Table 5.2 Mollusc species identified within the Stuarts Point aquifer. General depth of the mollusc fossil localities and their preferred habitats are also shown.124

Table 5.3 Common mollusc species and their environmental habitats identified in the sediments of the Stuarts Point aquifer. .................................................. 125

Table 5.4 Element means in the Stuarts Point aquifer sediments, as determined herein by XRF (n=36). ................................................................................ 129

Table 5.5 Mean Sb concentration for each of the proposed aquifer facies at Stuarts Point............................................................................................................. 141

List of Tables

O’Shea (2006) Page xvi

Table 5.6 Summary of eustatic changes in sea level and comparative depositional environments between the Bengal Delta Plain and the Stuarts Point aquifer.......................................................................................................... 154

Table 5.7 Potential arsenic sources to the Stuarts Point aquifer and the supporting/opposing evidence associated with each suggested source....... 163

Table 6.1 Elemental mean and ranges (mg kg-1) reported herein for the Stuarts Point aquifer sediments and their comparison to natural background concentrations for different matrices (full Stuarts Point data set provided in Appendix B1). ......................................................................................... 169

Table 6.2 Calculation of EF's for comparison to other coastal environments............. 171Table 6.3 EF Categories devised by Sutherland (2000). ............................................. 173Table 6.4 Pearsons correlations between solid phase arsenic and trace/major

element concentrations reported for the Stuarts Point geomorphic facies. Extremely strong correlations are highlighted in black, very strong correlations in grey, and strong correlations are bolded. Additional X-Y plots for selected correlations with arsenic are located in Appendix B9. ... 176

Table 6.5 Indicative amount of iron and aluminium present as oxides in the Stuarts Point matrix. ................................................................................................ 184

Table 7.1 Mean concentrations of general parameters and major ions in each groundwater type identified at Stuarts Point (mg L-1 unless stated). .......... 203

Table 7.2 Average chemical composition for each water group. Traces in ug L-1 and majors in mg L-1. ......................................................................................... 208

Table 7.3 Descriptive statistics for dissolved arsenic concentrations in each lithologically or hydrochemically dominated water group (ug L-1). ........... 209

Table 7.4 Proposed mobilisation processes (tested herein) from identified arsenic sinks at Stuarts Point and their influence on arsenic concentrations in groundwater................................................................................................. 212

Table 7.5 Mean and range for Stuarts Point LOI concentrations. Organic matter contents can be no higher than LOI............................................................. 226

O’Shea (2006) Page xvii

ABBREVIATIONS

statistical mean g L-1 micrograms per litre ‰ per mille ABS Australian Bureau of Statistics AHD Australian Height Datum (in metres) ANSTO Australian Nuclear Science and Technology Organisation APHA American Public Health Association ARMCANZ Agricultural Resource & Management Council of Australia & New

ZealandASS Acid Sulfate Soils BDP Bengal Delta Plain DDT dichlorodiphenyltrichloroethane (pesticide) DIPNR Department of Infrastructure, Planning and Natural Resources DL Detection Limit DLWC Department of Land and Water Conservation EC Electrical Conductivity EPA Environmental Protection Authority (E)XAFS (Extended) X-ray absorption fine structure spectroscopy GSA Grain Size Analysis HCA Hierarchial Cluster Analysis HFO Hydrous Ferric Oxide KMC K-Means Clustering LOI Loss on Ignition mbgs metres below ground surface mg L-1 milligrams per litre ML Multi-Level (piezometer) na not analysed nd not detected NHMRC National Health and Medical Research Council NSW New South Wales PC Principal Component PCA Principal Components Analysis ppm parts per million PVC polyvinyl chloride PWD Public Works Department SI Saturation Index TEM Transmission Electron Microscopy UNSW University of New South Wales WHO World Health Organisation WRC Water Resources Commission

O’Shea (2006) Page 1

1 INTRODUCTION

1.1 PROJECT INCEPTION

In 1995, the coastal Tomago Sandbed aquifer of eastern Australia underwent a

groundwater study to investigate the impact of heavy mineral sand mining on the

aquifer water quality (Coffey et al., 1996). A finding of this study concluded that

elevated levels of arsenic may be associated with the oxidation of marine pyrite during

lowering of the water table for mining associated activities. The presence of numerous

sand aquifers along coastal eastern Australia, which are stratigraphically similar to the

Tomago Sandbeds, raised concern over the quality of water presently being used for

domestic, stock, irrigation, town water and industrial activity along the eastern

Australian coastline.

To address these concerns, the New South Wales (NSW) State Government organised a

Taskforce comprising officers from the Environment Protection Authority (EPA);

Department of Infrastructure, Planning and Natural Resources (DIPNR – formerly the

Department of Land and Water Conservation, DLWC); Department of Minerals, Energy

and Health; the Hunter Water Corporation; and the Port Stephens Shire Council; to

investigate the occurrence of arsenic in coastal sand aquifers in NSW. Four hundred

and thirty three groundwater samples were collected from bores located within aquifers

spanning the entire length of the NSW coastline.

Thirty four samples returned arsenic concentrations above the Australian Drinking

Water Guidelines of 7 g L-1 As (NHMRC and ARMCANZ, 1996). Of these 34

samples reporting elevated arsenic, only three samples exceeded the World Health

Organisation (WHO) limit of 50 g L-1 As. The highest reported arsenic concentration

was 80 g L-1 from the town water supply bore at Stuarts Point, located on the mid-

North coast of NSW (Piscopo, 1996). Stuarts Point was thus given priority for further

investigation based on its use of groundwater for human consumption.

Chapter 1 - Introduction


Subsequently, in 1999 representatives from the University of New South Wales

(UNSW) visited Stuarts Point to inform residents of a regional groundwater study to be

implemented in the Stuarts Point sandy aquifer. This regional study focussed on the

following issues:

the impact of overextraction of groundwater for irrigation supplies;

the impact of Acid Sulfate Soils (ASS) and decreased water quality on

groundwater dependent ecosystems;

potential contamination issues arising from septic tanks, hillside runoff and

irrigation practices;

the regional impact of arsenic on the groundwater supply; and

potential for seawater intrusion into the freshwater aquifer following the re-

flooding of the Yarrahapinni Wetland upon opening of the tidal floodgates.

Results from this regional study can be found in Northey (2001) and Jankowski et al.

(2002). Seawater intrusion processes are discussed in Simpson (2004).

Above all else, the Stuarts Point community were especially concerned with the

elevated levels of arsenic present in their drinking water supply. Concurrently,

thousands of kilometres away in Bangladesh, news was emerging of what is now

referred to as the worst natural mass poisoning in history. Millions of Bangladeshi’s are

estimated to be at risk of arsenic poisoning from ingesting drinking water from an

aquifer with arsenic concentrations up to 200 times the WHO limit (Pearce, 1995). With

each passing moment, more people in Bangladesh were being diagnosed with arsenic

related illnesses, consequently causing social, economic and health implications for the

developing country. As time progressed, the occurrence of arsenic in groundwater of

other countries began to emerge. It seemed prudent to address the situation at Stuarts

Point as soon as possible. Thus, a study was established to examine solely, and in detail,

the occurrence of arsenic in the Stuarts Point groundwater environment. This detailed

study eventuated in a four-year research project. The results are provided within this

thesis.



1.2 RESEARCH AIMS In order to understand the occurrence and distribution of arsenic within the Stuarts Point

aquifer, three main aims have been deduced for this study:

1. To identify potential sources of arsenic in the aquifer, specifically determining

whether the arsenic is naturally elevated or anthropogenically influenced;

2. To delineate the current geochemical sinks of arsenic within the aquifer matrix;

and

3. To assess and determine the hydrogeochemical processes contributing to arsenic

transport and retardation (i.e., mobilisation) within the aquifer.

Establishing the source of arsenic within the aquifer matrix requires the detailed

examination of aquifer depositional conditions. Thus, a minor aim of this research is to

propose a thorough geomorphic evolution of the Stuarts Point aquifer using the data

generated during this study. Knowledge on local bedrock geology will also be updated.

1.3 HYPOTHESES

Addressing each of the three main aims, the following theories are hypothesised:

What is the Source of Arsenic in the Aquifer?

In the absence of any strong supporting evidence for anthropogenic contamination

sources in the study area, the presence of arsenic in the aquifer is hypothesised to be

naturally occurring. Four arsenic source theories are probable:

1. Arsenic has been contributed to the aquifer matrix via deposition of regionally

eroded geological units containing arsenic mineralisation;

2. Arsenic is derived from remnant seawater trapped in marine clay units deposited

during eustatic changes of sea level in the Quaternary (Smith et al., 2006);

3. The oxidation of arsenian pyrite present in ASS material contributes dissolved

arsenic to the groundwater (Smith et al., 2006); and/or

4. The underlying bedrock contains arsenic, which is being contributed to the aquifer

via upwards vertical leakage of groundwater (Smith et al., 2003).



The elevated concentrations of arsenic at Stuarts Point, potentially derived from sea

level inundation of the aquifer matrix during the Quaternary, is considered unlikely.

This notion is based upon the author’s belief that the commonly accepted model for

aquifer formation is hypothesised to apply to some, but not all, of the aquifer

depositional history. Therefore, arsenic is surmised to be present naturally in the aquifer

and is suspected to be a product of the weathering and re-deposition of (mineralised)

regional geology.

Pyrite oxidation from marine influenced ASS horizons in addition to some vertical

bedrock groundwater discharge, may contribute minor arsenic to the groundwater. Some

arsenic may have been contributed through anthropogenic activities but this is expected

to be minimal.

Where is the Arsenic Currently Being Stored in the Aquifer?

The current sinks of arsenic in the aquifer matrix are suspected to include sorption onto

oxide and clay mineral surfaces and precipitation as various solid phases. These

minerals may include iron and manganese oxides, marine clay minerals and/or iron

sulfides such as pyrite.

Which Geochemical Processes are Mobilising Arsenic into Groundwater?

The hydrogeochemical processes responsible for the mobilisation of arsenic into the

groundwater are expected to be the result of natural geochemical conditions. Chemical

heterogeneity may largely influence arsenic partitioning within solid phases, thus

affecting its occurrence in dissolved phase. Possible release mechanisms may include

desorption from and/or dissolution of arsenic-bearing oxide and clay minerals; and the

oxidation of minerals containing arsenic. Some mobilisation of arsenic may be

enhanced by anthropogenic activities resulting in disequilibrium of natural aquifer

conditions and further arsenic release.

1.4 SIGNIFICANCE OF THIS RESEARCH

Eustatic changes in sea level throughout the Quaternary led to the deposition of

numerous sand aquifers on the eastern Australian coastline, Stuarts Point being one of



them. The accepted model for these aquifer depositional conditions is that sediment was

supplied from off-shore during sea level stillstands. Thus, if the Stuarts Point aquifer

contains naturally elevated arsenic within aquifer sediments derived from off-shore, it is

possible that many other eastern Australian aquifers also harbour arsenic within their

matrices. If this is the case, aquifers worldwide that have been inundated by sea level

transgressions may be at risk of elevated arsenic occurrence. If this is not the case,

coastal aquifers with elevated arsenic occurrences should look elsewhere to establish the

arsenic source. The results of this study will thus contribute to risk assessment for

aquifers with similar depositional histories.

Stuarts Point is the first known coastal sand aquifer to exhibit elevated concentrations of

arsenic which are assumed to be geogenic. Identification of the current arsenic sinks

within the aquifer will lead to an increased scientific understanding of arsenic

partitioning within the solid phase of what is assumed to be a homogenous sand aquifer.

Detailed analysis of sediment-water interaction will increase scientific understanding of

arsenic mobilisation processes within this, and other, aquifer matrices.

The presence of elevated arsenic in this groundwater environment is also suspected to

influence other mediums. For example, arsenic may be accumulating in crops irrigated

with groundwater. This is suspected to vary according to crop type, the sporadic

occurrence of arsenic in the aquifer and the point of extraction of irrigation water from

the aquifer. In addition, groundwater discharging into the adjacent estuary is assumed to

be elevated in arsenic and therefore potentially accumulating in marine organisms

present in the estuary.

Finally, collection of a large amount of aquifer specific sedimentological,

hydrogeological and geochemical data will contribute to the overall management of the

aquifer by the NSW Department of Planning, Infrastructure and Resources, the Stuarts

Point town residents and the local farmers. This thesis is therefore expected to make a

significant contribution to science while providing practical information to aid in the

future management of this aquifer as a sustainable groundwater resource.



1.5 STUDY APPROACH

To investigate the main aims (and subsequently reject or accept the proposed

hypotheses) the tasks listed in Table 1.1 were executed. This list also differentiates

between existing knowledge previously available for incorporation in this investigation

and information that has been generated as part of this study.

1.6 THESIS STRUCTURE

The structure of this thesis is outlined in Figure 1.1. This introduction establishes a

summary of the problem to date and the need for its resolution. An overview of the site

location and geological knowledge available to date is provided in Chapter 2. The

literature review (Chapter 3) examines all aspects of arsenic chemistry relevant to its

occurrence in the environment and provides the reader with sufficient background

knowledge to critically assess the processes proposed herein. Chapter 4 outlines the

methods of analysis used within. An appreciable effort has been made to keep this thesis

clear and concise, with a strong focus on addressing the main aims, which form the bulk

sections of this thesis – Chapters 4 through 6 – covering arsenic sources, sinks and

mobilisation, respectively. Finally, Chapter 8 provides conclusions to the proposed

hypotheses, recommendations for the future management of the aquifer and any further

geochemical research that may be required. Supporting data are contained within the

appendices.

Tab

le 1

.1 T

asks

exe

cute

d in

this

thes

is a

nd n

ew d

ata

gene

rate

d as

a r

esul

t. Ta

skEx

isiti

ng In

form

atio

n D

ata

Gen

erat

ed H

erei

n D

eter

min

e re

lativ

e in

puts

(if

any)

from

ant

hrop

ogen

ic

arse

nic

sour

ces

His

toric

al re

cord

s –

spor

adic

and

in

com

plet

e; a

necd

otal

evi

denc

e A

sses

smen

t of a

ny re

cord

s pe

rtain

ing

to th

e us

e of

ars

enic

-co

ntai

ning

che

mic

als

in th

e vi

cini

ty o

f Stu

arts

Poi

nt –

liai

son

with

N

SW

Agr

icul

ture

, EP

A, L

and

& P

rope

rty In

form

atio

n, K

emps

ey

Shi

re C

ounc

il, lo

cal h

isto

rians

, Rut

ile Z

inc

Min

ing

Co.

& c

urre

nt

user

s of

the

land

E

stab

lish

pote

ntia

l for

ar

seni

c co

ntrib

utio

n fro

m

bedr

ock

geol

ogy

Lim

ited

and

cont

rast

ing

info

rmat

ion

avai

labl

e on

bed

rock

ge

olog

y; d

epth

and

lith

olog

y un

know

n

Ret

rieva

l of h

isto

rical

geo

logi

cal r

ecor

ds w

hich

hav

e be

en o

mitt

ed

from

the

mos

t rec

ent b

edro

ck g

eolo

gica

l stu

dies

; int

erse

ctio

n of

be

droc

k fro

m d

rillin

g pr

ovid

ed in

form

atio

n on

dep

th a

nd li

thol

ogy,

w

hich

sub

sequ

ently

sug

gest

ed a

faul

t exi

sts

in b

edro

ck w

hich

has

be

en o

verlo

oked

in re

cent

stu

dies

; exa

min

atio

n of

gro

undw

ater

co

mpo

sitio

n to

sup

port

or d

ism

iss

poss

ible

pre

senc

e of

faul

t E

stab

lish

pote

ntia

l for

ar

seni

c co

ntrib

utio

n fro

m

regi

onal

geo

logy

Det

aile

d ge

olog

ical

des

crip

tion

of

regi

onal

geo

logy

Reg

iona

l geo

mor

phic

mod

el

prop

osed

but

not

Stu

arts

Poi

nt

spec

ific

Lith

olog

ical

logg

ing

of a

quife

r sed

imen

ts to

dep

ths

grea

ter t

han

ever

sam

pled

bef

ore;

sed

imen

tolo

gica

l met

hods

and

sta

tistic

al

anal

ysis

of s

edim

ent c

hem

istry

ena

ble

aco

mpr

ehen

sive

geo

mor

phic

mod

el to

be

deve

lope

d fo

r dep

ositi

on

of th

e S

tuar

ts P

oint

aqu

ifer

Exa

min

atio

n of

cur

rent

ar

seni

c si

nks

in th

e aq

uife

r m

atrix

Nil

X-ra

y D

iffra

ctio

n; S

cann

ing

Ele

ctro

n M

icro

prob

e; S

X50

Ele

ctro

n M

icro

prob

e; S

eque

ntia

l Ext

ract

ions

; X-ra

y Fl

uore

scen

ce; S

tatis

tical

A

naly

ses;

bor

ehol

e pr

ofile

s A

sses

smen

t of g

roun

dwat

er

cond

ition

s su

itabl

e fo

r the

m

obili

satio

n of

ars

enic

Pre

limin

ary

stud

ies

sugg

estin

g ar

seni

c m

obili

satio

n pr

oces

ses

are

geoc

hem

ical

ly c

ompl

ex

Sta

tistic

al a

naly

sis

of m

ore

than

200

gro

undw

ater

sam

ples

from

10

mul

ti-le

vel p

iezo

met

ers

with

met

re d

epth

incr

emen

t sam

plin

g po

ints

; ana

lysi

s of

geo

chem

ical

cor

rela

tions

bet

wee

n ar

seni

c an

d m

ore

than

50

elem

ents

in a

queo

us p

hase

; ass

essm

ent o

f ge

oche

mic

al s

uita

bilit

y fo

r rel

ease

from

iden

tifie

d ar

seni

c si

nks

in

the

aqui

fer;

geoc

hem

ical

mod

ellin

g to

sup

port

prop

osed

m

obili

satio

n pr

oces

ses;

ass

essm

ent o

f cur

rent

pro

cess

es a

ffect

ing

arse

nic

pres

ence

and

mob

ility

in th

e aq

uife

r In

fluen

ce o

f ars

enic

in th

e su

rroun

ding

env

ironm

ent

Nil

Indi

catio

n of

ars

enic

acc

umul

atio

n in

cro

ps; a

nd in

fluen

ce o

f ar

seni

c in

the

adja

cent

est

uary

Figure 1.1 Thesis Structure


2 THE STUARTS POINT STUDY AREA

2.1 LOCATION AND ENVIRONMENTAL CONDITIONS1

Stuarts Point is located on the NSW mid-north coast (Figure 2.1), approximately 400

km north of Sydney, Australia (lat. 30º S, long. 153º E). The region is exposed to a

temperate climate with average maximum summer temperatures of 27º C in February

and average minimum winter temperatures of 11º C in July (Bureau of Meterology,

2005). Higher rainfall is received in summer than in winter, with maximum March

rainfall (average 194 mm - Bureau of Meteorology, 2005) generating excellent

conditions for extensive groundwater recharge via these increased precipitation rates.

Climate can vary, however, as observed during the course of this four year investigation

which saw floods, El Nino-induced bushfires and extreme drought plague much of the

region.

2.2 THE MACLEAY RIVER CATCHMENT

The Stuarts Point aquifer is positioned on the northeastern margin of the Macleay River

catchment, which covers approximately 11,450 km2. The catchment consists mainly of

floodplain deposits downstream of Kempsey, the region’s largest town (Figure 2.2).

Upstream of Kempsey, the rocky terrain of the New England Fold Belt characterises the

regional landscape.

The dominance of the floodplain in the surrounding region keeps topographic gradients

at Stuarts Point to a minimum, resulting in heights generally <5 m Australian Height

Datum (AHD). The Yarrahapinni Mountain rises approximately 495 m to the immediate

northwest of the Stuarts Point township (Figure 2.1).

1 A selection of site photographs can be viewed in Appendix C.

Figure 2.1 Location of the Stuarts Point sand aquifer (contained within the dotted area). Multi-level piezometers used in this study are also shown.

Figu

re h

as b

een

rem

oved

due

to C

opyr

ight

Agr

eem

ents

Figu

re 2

.2

Bou

ndar

ies

of t

he M

acle

ay R

iver

Cat

chm

ent.

Stua

rts

Poin

t is

loc

ated

on

the

nort

heas

tern

mar

gin

of t

he c

atch

men

t w

here

flood

plai

n de

posi

ts d

omin

ate

(NSW

EPA

, 200

5).

Chapter 2 – Study Area


The lower reaches of the mountain are used for banana, strawberry and macadamia

plantations while the upper reaches consist of wet sclerophyll forest. Much of the low-

lying land to the south of the township consists of heathland, sporadic residential lots

and agricultural crops including potatoes, stone fruits and avocadoes. The Macleay

River estuary located adjacent to Stuarts Point is used extensively for recreational

activities and commercial fishing, prawning and oyster production.

2.2.1 Regional Geology Regionally, the Macleay River drains the hinterland of the New England Fold Belt

(NEFB). Geologically, Stuarts Point lies on the eastern border of the southern section of

the NEFB (Figure 2.3).

Figure 2.3 Extent of the New England Fold Belt and its adjacent basin margins (modified from Palfreyman, 1984 and Leitch, 1974).

The NEFB is a complex system of Palaeozoic-Triassic sediments, intruded by acid to

basic volcanics, moderately to strongly folded and faulted, with metamorphism ranging

from very low to high grade (Palfreyman, 1984). Mineralisation is associated with

volcanic intrusions, fault zones, placer deposits and overlying Tertiary basalt. Parts of

the basin have major breaks in the stratigraphic record, either due to erosional periods or



lack of detailed geological mapping. Over the past century numerous geologists have

debated the geological history of the NEFB and its associated complexities. Postulated

theories and supporting evidence can be found in reviews by (chronologically)

Woolnough (1911); Kenny (1935); Voisey (1958); Lindsay (1969); Packham (1969);

Campbell (1969); Leitch (1974) and Gilligan and Barnes (1990)2. Figure 2.4 shows the

present day geological subdivision of the southern NEFB. Each subdivision is a fault-

bounded block representing an interval of deposition and/or deformation.

Figure 2.4 Geological subdivision of the southern NEFB. Stuarts Point is situated on the Nambucca Block (modified from Gilligan et al., 1992).

2 N.B. This list is not conclusive.



Table 2.1 provides a brief geological history that has been inferred from the above

references for the southern section of the NEFB.

Table 2.1 Geological History of the NEFB. Cambrian – Carboniferous

On the western margin of the southern NEFB a volcanic chain provided the major source of detritus for the fold belt (Leitch, 1974). Parts of this chain are now presumed to be buried underneath the Sydney Basin. Associated with this volcanic chain were fore-arc basins, now represented by the Tamworth Belt and the Hastings Block, which accumulated metasediments from these distal volcanic sources (Gilligan et al., 1992). Further to the east in the region of the Central, Dyamberin, Coffs Harbour, Port Macquarie and Nambucca Blocks, sediments from the deep ocean floor accumulated as accretionary prisms – sediments derived from off-scrapes from the top of the subducting plate.

An overall change from andesetic volcanism in the Devonian to basalt-andesite-rhyolite in the Early Permian is evident in both exposed rocks of the chain, and detritus in the sedimentary material (Leitch, 1974). This sedimentary material deposited in the fore-arc basins consisted largely of mudstones and greywackes with very little quartz yet abundant in volcanic rock fragments (Voisey and Packham, 1969). Therefore, volcanism provided much of the sedimentary material, with minor contributions from terrestrial sources, during this time.

Deep sea sediments deposited in the accretionary prism have been dated Lower-Palaeozoic chert-jasper-siliceous argillite-basalts and are suggested remnants of oceanic crust (Leitch, 1974). Sediments of similar lithology (and hence depositional environment) occur throughout many of the Palaeozoic dated units.

Late Carboniferous – Permian This was the major period of deformation and uplift that transformed the accumulated metasediments into the NEFB (Gilligan et al., 1992). The late Carboniferous saw the progression of the shoreline eastward via deposition of volcanic detritus coupled with terrestrial slumping and glacial sediments.

Reworking of basement rocks and deposition of volcanics continued into the Permian, however large scale displacement around this time prevents a good succession in the stratigraphic record.

TriassicMassive discordant granitoids intruded the NEFB in the mid-Triassic. There are two main granitoid belts – the Gundle belt and the coastal belt. These intrusions generally possess sharply defined contacts and well delineated metamorphic aureoles. A full mineralogical description of the intrusion within the study area can be found in Section 2.3.2.2.



2.2.2 The Macleay Fluvial-Deltaic Floodplain

The floodplain consists of a variety of soils and sediments deposited during eustatic

changes in sea level during the Quarternary. Sediments consist of alluvial terraces

(Walker, 1970), back barrier swamp deposits, beach barrier sands (Hails, 1968),

estuarine muds and silty clay sequences (Eddie, 2000). Soil development is highly

dependent on sediment type and can vary from sandy podzolized soils to organic

topsoils with high ASS potential (Haskins et al., 2000). Aboriginal shell middens have

been identified in the area and are considered to be of archaeological significance

(Sullivan and Hughes, 1978).

2.3 DEPOSITION OF THE STUARTS POINT AQUIFER

The coastline of NSW is characterised by zeta-curved sandy bays existing between

rocky headlands and Stuarts Point is no exception to this description. A sandy barrier

beach extends from Grassy Head at the northern end of the aquifer to the Macleay River

in the south. Numerous geomorphic investigations have been conducted along the NSW

coast (Langford-Smith and Thom, 1969; Roy and Thom, 1981; Thom, 2003) and more

specifically the mid-north coast (Voisey, 1934; Hails, 1964, 1967, 1968; Walker, 1970).

The Stuarts Point sandy aquifer boundary is shown in Figure 2.1. Sediments located

adjacent to this main water-bearing unit consist of backswamp deposits and estuarine

clays; including the Yarrahapinni wetland. A full description can be found in Eddie

(2000). These sediment deposits have not been investigated as part of this study,

information on these can be found in regional studies conducted by Smith et al. (2003)

and Northey (2001).

2.3.1 Previous Geomorphic Models The most common geomorphic model applied to the northern coast of NSW is the

development of inner (Pleistocene) and outer (Holocene) beach barrier systems. Roy

and Thom (1981) provide a generalised facies relationship for dual barriers on the

central NSW coast (Figure 2.5). Despite the numerous sediment lithologies and past

shorelines of the Macleay River floodplain (Hails, 1968; Walker, 1970; Eddie, 2000),



the aquifer investigated herein focuses on a sandy water-bearing unit generally

described as a coastal barrier. It is this stratigraphy which is commonly adopted by

investigators in the area (Eddie, 2000; Smith et al., 2003; 2006).

Figure has been removed due to Copyright Agreements

Figure 2.5 Generalised section of a dual barrier system (Roy and Thom, 1981).

Hoyt (1966) proposes the dual barrier system primarily develops due to submergence of

dune or beach ridges adjacent to pre existing shorelines; that is, sediment is derived

from off-shore and is pushed on-shore during stillstands. Prior to this deposition,

however, existed a pre-Pleistocene shoreline present inland of the current coastline. This

former shoreline is illustrated nicely in Hails’ (1968) existing geomorphic model for the

region shown in Figure 2.6. This figure shows the former Pleistocene shoreline and

headlands at Hat Head and Smoky Cape were once offshore islands and have since been

joined to the coastline by deposition of a dual barrier system.




Figure 2.6 Pre-Pleistocene and current Holocene coastlines on the NSW mid-north coast (Hails, 1968).

Eddie (2000) classified the Stuarts Point aquifer as an inner barrier unconsolidated

siliceous beach and dune sand, denoted by ‘sp’ on the extract of his soil landscape map

presented in (Figure 2.7).




Figure 2.7 The Stuarts Point aquifer (outlined) denoted by Eddie (2000) as an inner barrier beach/dune sand – ‘sp’ (extracted and modified from Eddie, 2000).

Hails (1968) described the area simply as ‘abandoned beach ridges’ in contrast to actual

inner and outer barrier systems. He does state that deflection of the Macleay River most

probably occurred during formation of the outer barrier which is thus assumed to be

positioned on the eastern (seaward) side of the Macleay Arm and thus not within the

limits of the Stuarts Point aquifer investigated herein. The outer Holocene barrier is also

quartzose and contains abundant shell material. The outer Holocene barrier has formed

in the last 7,000 years during which time there has been a constant and plentiful onshore

supply of sand (Hails, 1968).

In terms of groundwater environments elevated in dissolved arsenic, a quartz-rich sand

barrier deposit constructed from sediment supplied from offshore proved to be a unique

and alarming occurrence of such natural contamination; thus sparking the need for this

research. This commonly accepted dual barrier geomorphic model for the deposition of

the Stuarts Point aquifer will be assessed in Chapter 5 – Original Source of Arsenic to

the Aquifer.



2.3.2 Bedrock Geology

2.3.2.1 Lithology and Structure The underlying bedrock has been deduced as belonging to the NEFB, however, a review

of previous literature relating to the Stuarts Point (site specific) geology indicates

inconsistent theories on the likely bedrock lithology present beneath the unconsolidated

aquifer. Kenny (1935) explains the likely cause of these inconsistencies in his

geological reconnaissance of the north coastal region.

“In common with neighbouring tracts of country practically the whole of the

area traversed is heavily timbered. Rock outcrops are masked by a cover of scrub,

grass, and waste material to the extent that favourable exposures are restricted

almost wholly to sections along the sea-coast to road and railway cuttings, to

gullies in the upper reaches of stream systems, and to mine workings. For these

reasons it has been found impracticable to trace geological boundaries save in a

very general sense and no attempt at any detailed interpretation of structural

relationships and stratigraphical succession has been made.” (pg 85)

The inaccessible terrain surrounding Stuarts Point has therefore made ground truthing

and subsequent geological mapping difficult. Geological boundaries presented herein

may have been inferred from the limited data available and should thus be interpreted

with caution.

Figure 2.8 shows several possible bedrock lithologies proposed by various geologists.

The earliest account was conducted by Voisey (1950). A possible fault, termed the

‘Kempsey Area Fault’, was indicated to extend from the bedrock outcrops west of

Stuarts Point, trending in an easterly direction towards the coast and dissecting the

Stuarts Point study area. The Kempsey Area Fault was hypothesised to dissect northern

lower Palaeozoic slates and phyllites from southern upper Palaeozoic micaceous

mudstones of the Warbro (Permian) stage of the Macleay series. According to Voisey

(1950), these upper Palaeozoic sediments have been lowered relative to the northern

lower Palaeozoic sediments; and were already folded when the Kempsey Area Fault

occurred late in the Upper Palaeozoic orogeny.

Figure 2.8 Various bedrock lithologies proposed for Stuarts Point. The Stuarts Point aquifer is contained within the dotted area.



Voisey (1958) again suggested the presence of the Kempsey Area Fault, described as a

transcurrent fault, once again separating Permian sediments in the south from lower

Palaeozoic sediments in the north. Several years later Voisey and Packham (1969)

provided a review of the New England Region. The same theories were postulated

except that the lower Palaeozoic sediments north of the Kempsey Area Fault were re-

named the Nambucca Phyllites.

Suppel (1974) made no mention of the Kempsey Area Fault. Figure 2.8 marks the

boundary proposed by Suppel for the Nambucca Beds, consisting of slate, phyllite,

schistose, sandstone and schistose conglomerate of probable Permian age. No reason

was given as to the difference in proposed boundaries for the Nambucca Beds

previously described by Voisey and Packham (1969). It is thought that the boundary

was simply inferred, but could not be clearly identified due to the presence of alluvium

covering accessible outcrops.

Similarly, Gilligan et al. (1992) provided a metallogenic map of the Nambucca Block

without any indication of a fault occurring in the Stuarts Point area. The area where

Suppel’s (1974) Nambucca Beds were postulated to occur were divided into Nambucca

(Pee Dee) and Kempsey Beds by Gilligan et al. (1992). The Kempsey Beds were in turn

described as consisting of lithic sandstone, mudstone, pebbly sandstone and minor

conglomerate of late Carboniferous to early Permian age and most probably correspond

to Leitch’s (1974) Kempsey Beds. Leitch (1974) had previously provided an account of

regional geology in the Nambucca Block. While no approximate boundaries were

inferred for the Stuarts Point area, the Kempsey Beds were thought to overlie the

Nambucca (Pee Dee) Beds in the southeast of the Nambucca Block (i.e., the northwest

corner of Stuarts Point). Leitch (1974) described the Kempsey Beds as consisting of

interbedded lithic sandstone and siltstone with occasional conglomerate horizons. Both

sandstones and siltstones were indicated to potentially be pyritic; many of them showing

evidence of organic reworking. The Pee Dee Beds were described by Leitch (1974) to

be dominated by cleaved siltstone. All rocks were deposited in a marine environment

with terrigenous sediment contributions.



In their review of the geology of Stuarts Point, Smith et al. (2003) concluded from the

above evidence that the unconsolidated aquifer sediments must overlie the Kempsey

Beds. They also stated that the upper reaches of the catchment were characterised by the

early Permian Pee Dee Beds of slaty siltstone, lithic sandstone and minor diamictite.

Gilligan et al.’s (1992) boundary for the Pee Dee Beds does not extend as far southeast

to the Stuarts Point aquifer, however drilling evidence obtained herein, in addition to the

location of a slate/phyllite quarry at Stuarts Point, indicates the Nambucca (Pee Dee)

Beds may in fact underlie the northern part of the Stuarts Point aquifer.

Smith et al. (2003), Gilligan et al. (1992) and Suppel (1974) made no reference to

Voisey’s previous papers (1950; 1958) in which the Kempsey Area Fault was listed as a

principal structural feature. It seems that the possibility of this Kempsey Area Fault

being present has been neglected in the most recent geological literature without any

plausible explanation of its omission. The terrain surrounding Stuarts Point has been

noted by many authors as being harsh and inaccessible. If the fault is present, the

unconsolidated deposits of the Stuarts Point aquifer cover it. Combined with the

inaccessible terrain, this fault may exist but to date has not been accurately mapped. In

addition, many of the later geological investigations were conducted on a regional scale

rather than a site specific scale and therefore attention to an inconspicuous fault at

Stuarts Point may have been overlooked.

2.3.2.2 MineralisationThe Nambucca Block contains a coastal granitoid belt extending approximately 40 km

in length and 10 km wide and comprising small discrete granitic intrusions (Gilligan et

al., 1992). In the vicinity of Stuarts Point an adamellite granitoid intrusion forms Mt

Yarrahapinni in the northwest of the study area (Figure 2.9).

Substantial mineralisation occurs within the granitic intrusion and contact aureole

forming Mt Yarrahapinni. The intrusion and associated aureole span approximately 1.5

km by 2.5 km and are located to the southwest of a larger intrusion (Suppel and Hobbs,

1977). Mineralisation is directly associated with the granitic intrusions which are

located in complexly deformed low grade metasediments of the Nambucca Block

(Suppel and Hobbs, 1977). Deposits are present in both the granitic and



metasedimentary rocks – the phyllites and slates of the Nambucca (Pee Dee) Beds.

Further description of mineralised deposits are discussed below and are largely

summarised from Gilligan et al. (1992).


Figure 2.9 Coastal granitoid belt forming Mt Yarrahapinni (modified from Gilligan et al., 1992).

Molybdenite is sparsely disseminated throughout the adamellite and therefore is the

dominant exploited form of mineralisation present at Mt Yarrahapinni, with three

deposits having previously been mined. Deposits of tungsten, silver, base metals (Pb,

Zn) and gold occur in and around the Yarrahapinni adamellite. Mineralisation grades

from an outward zoning molybdenum zone (I), to a silver-lead zone (II) and finally to a

silver-arsenic zone (III). On the Stuarts Point side of Mt Yarrahapinni these mapped

zones are incomplete, indicating the lack of accurate geological mapping specific to the

area (Figure 2.10).




Figure 2.10 Zones of mineralisation surrounding Mt Yarrahapinni (modified from Gilligan et al., 1992). Numbers refer to recognised mineral deposits described in Gilligan et al. (1992). 772-774 are disseminated/vein Mo deposits.

The exact extent of the intrusion remains uncertain. Air borne geophysics was thus

obtained to aid in the interpretation of the extent of the adamellite intrusion forming Mt

Yarrahapinni. However, the quality of this data was poor and very little processing

could be conducted. An anomaly was detected in the vicinity of the Stuarts Point

aquifer, however, the poor data quality prevented further analysis or identification of the

anomaly. It should be assumed that the granitoid intrusion may extend underneath the

Stuarts Point aquifer further than previously mapped by Gilligan et al. (1992).

Mineralisation occurs primarily as (Gilligan et al., 1992):

disseminated - particularly Mo with associated chalcopyrite, pyrite and pyrrhotite;

joint-controlled sheeted veins - quartz-molybdenite-scheelite veinlets hosted by the

adamellite;



fracture-controlled quartz veins - discrete molybdenite-quartz veins within the

adamellite and associated hornfels, with disseminated or sporadically associated Ag-

As-Zn sulfides; and

fracture-controlled siliceous sulfide veins - joint controlled parallel vein system of

high-grade Ag-Pb-Zn mineralisation hosted by the hornfels.

In a geochemical survey of the Yarrahapinni - Way Way State Forest area, Suppel and

Hobbs (1977) found arsenic stream sediment concentrations were high, widespread and

frequently associated with anomalous molybdenum and copper zones. They further

state,

“the results indicate that arsenic is associated with the granitic intrusions” (pg

43).

In addition, Kenny (1935) noted,

“narrow veins of quartzose material, together with sulfides of lead, zinc, iron and

arsenic, the last-mentioned predominating, are developed along joints in

claystones close to the granite contact” (pg 86).

Arsenic is therefore undoubtedly associated with rocks located in close proximity to and

potentially underneath the Stuarts Point aquifer.

2.3.3 Hydrogeology The township extracts groundwater from a designated borefield to supply its 781

residents (Australian Bureau of Statistics, 2001) in addition to localised extraction for

agricultural irrigation. Domestic groundwater is treated to remove arsenic, however

most irrigation water is extracted via on-farm spearpoint bores and applied to crops

untreated.

Groundwater levels were measured in May 2001 and indicate a general flow direction

from Mt Yarrahapinni in the northwest, flowing south-southeast through the sand

aquifer and discharging into the Macleay Arm (Figure 2.11).



Figure 2.11 Groundwater contour map (mAHD) for the Stuarts Point sand aquifer.

Extensive aquifer testing is described in Northey (2001) where two interpretative

methods for hydraulic conductivity provided a range of 0.01 to 35.73 m/day. This

conductivity range includes values obtained for the adjacent clayey swamp deposits,

which produced low hydraulic conductivity values. The barrier sands were reported to

exhibit the highest hydraulic conductivity, with overall values highly dependent upon

aquifer heterogeneity. Actual hydraulic conductivity values for the sample locations

described herein are unknown, however it is assumed that the clayey units have

conductivities closer to 0.01 m/day, while groundwater moving through the sand layers

are expected to be close to 35 m/day.



Recharge to the aquifer is assumed to be dominated by diffuse precipitation and

potential recharge from bedrock groundwaters below the sandy aquifer. Isotopic

investigations conducted by Smith et al. (2003) suggest Yarrahapinni bedrock

groundwaters have endured a process of isotopic fractionation due to orographic

precipitation and the altitude effect described by Bortolami et al. (1979). They thus

exhibit 18O and 2H values of –5.2 and –26.3‰ respectively. In contrast, previous

work by Smith et al. (2001) appear to have mistakenly suggested that recharge

groundwater associated with the Yarrahapinni fractured (bed)rock system exhibits 18O

and 2H values up to approximately –7.3 and –44.2‰ respectively. Northey (2001)

graphed the same data according to lithological units in the study area and found that

Holocene sands plotted in this latter region (as opposed to bedrock waters only) and that

Holocene/Pleistocene sands and alluvial clays all plotted within the first stated range.

Additionally, the presence of the 500m high Mt Yarrahapinni is considered too low to

be affected by the altitude effect. For these reasons, the isotopic data is deemed

unreliable in the interpretation of bedrock discharge to the overlying unconsolidated

aquifer. Further isotopic investigations are known to be underway in the aquifer, but at

the time of writing this thesis these results had not yet been made publicly available.

The assumed discharge areas are the adjacent wetlands and Macleay River Arm estuary.

Water level data loggers examined by Smith et al. (2001) show the sandy aquifer

responds to rainfall events with a short time lag evident between precipitation events

and an observed rise in groundwater levels. Given the sandy nature of the aquifer

matrix, short response times to groundwater recharge can be expected. The close

proximity of the aquifer to the coastline may also induce diurnal variations in

groundwater levels influenced by the daily movement of tides. During extreme weather

conditions, such as floods and drought, both of which were observed during the course

of this investigation, the aquifer may respond accordingly. Indeed, a decline in

groundwater levels of approximately 1 m was observed between winter and summer. In

the summer, increased groundwater extraction for town and irrigation supplies may

provide a change in the flow conditions observed in Figure 2.11. While temporal

changes were not investigated during this study, it is important to realise that a change

in groundwater flow conditions will influence solute transport pathways. Bearing this in



mind, the interpretations presented herein represent solute transport occurring under the

conditions observed during the sampling period.

2.4 PREVIOUS INVESTIGATIONS

Piscopo’s (1996) Taskforce report provided a baseline investigation for arsenic

occurrence within the Stuarts Point aquifer. Piscopo proposed there was sufficient

evidence to suggest that elevated levels of arsenic along the coast were:

localised, yet variable;

associated with disturbed coastal dune sands where pyrite was present;

below irrigation and stock guidelines;

predominantly below 50 g L-1 As;

potentially linked to acid sulfate soils; yet

insufficient evidence was found to determine a relationship between groundwater

abstraction and elevated arsenic concentrations.

It also raised further questions. Why was arsenic elevated at Stuarts Point and not in

other coastal aquifers? What is the definite source of the arsenic? Who, if anyone, is to

blame? How is it getting into the groundwater, and where is it going? What can we do

to minimise, or avoid, further arsenic contamination in this and other aquifers?

The horizontal distribution of arsenic at Stuarts Point was investigated by Smith et al.

(2000). This investigation sampled groundwater from wells installed in nearby Mt

Yarrahapinni (assumed bedrock), the inner Pleistocene Barrier, clay and swamp

deposits adjacent to the Stuarts Point sandy aquifer and the aquifer itself (Figure 2.12).

Arsenic was reported up to 70 g L-1 in wells installed at various depths and in different

stratigraphic facies, indicating a non-localised and sporadic occurrence of arsenic in the

local environment. Follow-up vertical studies of As distribution in the high yielding

sandy aquifer reported concentrations of up to 330 g L-1 (Smith et al., 2003).




Figure 2.12 Regional arsenic distribution reported by Smith et al. (2000). The Stuarts Point aquifer and multi-level piezometers installed for this study are marked.

Given the commonly accepted geomorphic model of barrier sand aquifer stratigraphy,

Smith et al. (2006) deduced that a regional arsenic source was not likely. Instead, they

proposed that the source of arsenic originated in the following manner,

“during sea-level fluctuation in the late Pleistocene and Holocene, sorption of

arsenic into low-permeability clays occurred”

which indicates that it was the natural concentrations of arsenic present in the

transgressing sea that were responsible for its original presence in the aquifer. This

theory, if proved correct, could have ramifications for many coastal aquifers subjected

to sea level transgressions, not solely in Australia, but potentially worldwide.



Smith et al. (2003) also examined the aquifer groundwater composition and suggested

two principle geochemical processes contributing to arsenic mobilisation in the aquifer.

Arsenic reached concentrations of 85 g L-1 at a depth of 10-11 m and was suggested to

be

“present due to either the dissolution of Al hydroxides and release of adsorbed

As(V)… or concurrently via pH-influenced desorption of As enriched Fe oxides”

(pg 1490).

The greatest arsenic peak (337 g L-1) occurred at a depth of 25 m and was proposed to

be derived from

“leaching of the aquifer’s sandy-clayey matrix by groundwaters rich in HCO3-

and of high pH” (pg 1493).

More thorough investigations continued and included speciation of arsenic in aqueous

phase, but the method used was later found to be questionable during the course of the

investigations herein. Limited solid phase data were available to accurately determine

the source or sink of arsenic in the aquifer matrix. Therefore, assumptions governing

arsenic mobility could only be postulated based on the available aqueous data, with the

conclusions suggesting several plausible geochemical processes were governing arsenic

mobility in the aquifer. No definite arsenic source, sink or dominating mobilisation

process could be identified and sufficiently supported with the limited data available at

that time. The culmination of Smith’s research (Smith, 2005) recommends further

examination is required to better understand the arsenic geochemical cycle at Stuarts

Point. This investigation aims to fill that gap.

To date, no further published material has been found relating to the Stuarts Point

hydrogeochemical environment.


3 ARSENIC LITERATURE REVIEW

In the last decade there has been a dramatic increase in the number of studies

investigating arsenic in the environment. Much of this can be attributed to the discovery

of elevated levels of arsenic in domestic groundwater supplies throughout the world,

with particular respect to the arsenic crisis in Bangladesh. The severity of the problem

in Bangladesh has led to an increased amount of international scientific research being

conducted on arsenic toxicity, mobility, occurrence, chemical behaviour, cycling and

bioaccumulation. As such, a number of recent reviews on arsenic in the environment

have been provided by Welch et al. (2000); Smedley and Kinniburgh (2002); Welch and

Stollenwerk (2003); and Naidu et al. (2006); in addition to an entire Applied

Geochemistry journal volume (Volume 18, 2003) devoted solely to arsenic

geochemistry. Further to these recent reviews, earlier publications by Ferguson and

Gavis (1972) and Cullen and Reimer (1989) provide a solid basis on the theoretical

chemistry of arsenic in the environment.

Given the amount of published literature on this current affair, this review will focus on

topics considered by the author as appropriate background information on arsenic issues

most relevant to this study. Where possible, references for further reading will be

provided on elements that are less critical to this investigation and are therefore only

briefly discussed herein.

3.1 GENERAL PROPERTIES OF ARSENIC

3.1.1 Periodicity Arsenic belongs to Group 15 of the Periodic Table which comprises N, P, As, Sb and

Bi. The properties of these elements vary as the group is descended, with N and P

classified as non-metals; As and Sb as metalloids; and Bi being the only metal present

within the group. Group 15 is one of the few groups to exhibit the full range of element

chemistries – from non-metal to metalloid to metal – in a single group (Carmalt and

Norman, 1998).

Chapter 3 – Arsenic Literature Review


Arsenic is considered a metalloid (semi-metal) because it resembles a metal in

appearance but behaves like a non-metal. Other metalloids include B, Si, Ge, Sb, Se and

Te. These elements conduct electricity, but less effectively than metals, and are

therefore often referred to as semi-conductors (Moeller et al., 1989). Arsenic is

monoisotopic – it has only one stable isotope, 75As. The common properties of arsenic

are listed in Table 3.1.

Table 3.1 Common Physico-Chemical Properties of Arsenic (adapted from Carmalt and Norman, 1998; Moeller et al., 1989; Reimann and de Caritat, 1998).

Atomic Number 33 Atomic Mass 74.92 Melting Point 814 CBoiling Point 610 C

Main Oxidation States -3, +3, +5, 0 Density 5.73 g/cm3 (grey)

1.97 g/cm3 (yellow) Isotope 75 As (100% abundance)

Electronegativity 2.0 Atomic Radius 133 pm

Arsenic exists in the environment in the following oxidation states:

arsenate (As 5+)

arsenite (As 3+)

arsenic (As 0)

arsine (As 3-)

with oxyanions of arsenate and arsenite most common in aqueous environments. Both

oxyanion states form stable bonds with carbon, resulting in the formation of organo-

arsenic compounds. Organo-arsenic forms may be produced from biological activity but

are rarely present in sufficient amounts, unless derived from anthropogenic

contamination sources. Arsenic has a high gaseous mobility in the atmosphere (as toxic

arsine gas – As3-) due to its low boiling point (Langmuir, 1997). The free elemental

state (As0) occurs only rarely.



3.1.2 Toxicity Arsenic is considered an essential element and is routinely found in the body, however

many arsenic compounds are toxic. Akin to heavy metals, the toxicity of arsenic largely

depends on its chemical form, with toxicity decreasing in the order: arsine > arsenite >

arsenate > monomethylarsonic acid (MMA) > dimethylarsinic acid (DMA) (Jain and

Ali, 2000).

Arsine gas is the most toxic compound and if readily inhaled may cause lung cancer

(Bissen and Frimmel, 2003). Arsenite is resorbed faster into biological systems than

arsenate and hence is considered to be 60 times more toxic (Ferguson and Gavis, 1972).

Both oxidation states inhibit the energy-linked functions of the mitochondria (Bissen

and Frimmel, 2003). In addition, arsenic compounds block cell and tissue respiration,

can paralyse smooth muscle and cause numerous small haemorrhages (DMR, 1995).

Organo- arsenic compounds are least toxic to humans because the methylation of

arsenic reduces the affinity of the compound for tissue (Vahter and Marafante, 1988).

Arsenic is carcinogenic and may cause lung, bladder, liver, renal and skin cancers

(Bissen and Frimmel, 2003). A dose of 100 mg As2O3 can cause chronic arsenic

poisoning with 130 mg As2O3 considered fatal (DMR, 1995). Chronic symptoms

include fatigue, gastritis, anorexia and hair loss; with long term exposure producing

hyperkeratosis (thickening of the skin), cardiovascular diseases, disturbances in the

peripheral vascular and nervous systems, circulatory disorders and liver and kidney

disorders. Arsenic is deposited in hair, nail, bones and skin; and is readily excreted

through urine after three to four days. The body can detoxify (to an extent) arsenic

compounds by methylation to reduce its affinity for human tissue (Pontius et al., 1994).

Humans accustomed to arsenic ingestion have been shown to withstand greater amounts

of arsenic intake than those who are not accustomed to its presence within the body

(Morton and Dunnette, 1994). There is an increased amount of research being

conducted on the epidemiology of arsenic (Aguirre-Banuelos et al., 2004; Lu et al.,

2004) with particular reference to spatial distribution between arsenic health impacts

and consumption of arsenic-rich groundwater (Mandal and Biswas, 2004; Rahman et

al., 2005).



3.1.3 Arsenic Production and Use Arsenic has long been used for its toxic and medicinal properties (Table 3.2). Arsenical

compounds were used in the Orient 2,000–3,000 years ago. Therapeutic uses continued

into the nineteenth century when classic examples of chemotherapy originated from the

use of arsenical compounds (Ferguson and Gavis, 1972). Diseases such as asthma,

malaria, diabetes and tuberculosis were treated with arsenic until the discovery of

antibiotics. In World War I and the Vietnam War, organically bound arsenic (such as

‘agent blue’) was used for chemical warfare purposes (Bissen and Frimmel, 2003).

Table 3.2 Past and present uses of arsenical compounds (adapted from Nriagu and Azcue, 1990; Bissen and Frimmel, 2003; WHO, 1992).

Industry UsesAgriculture Pesiticides, herbicides, insecticides, fungicides, defoliation Medicinal Early chemotherapeutic treatments, antisyphilitic drugs, treatment

of diabetes, malaria, amebiasis, sleeping sickness, embalming of corpses

Livestock Animal feed additives, cattle and sheep dipping, algaecides, disease prevention (heartworm infection, swine dysentery)

Metallurgy Alloys, hardening agents in battery plates Electronics Semi-conductor applications, solar cells, light emitting diodes

(digital watches), optoelectronic devices Industry Catalysts, dyes and soaps, ceramics, antifouling paints,

pyrotechnics, electrophotography, glassware, pharmaceutical substances, dehairing skins in tanning

Prior to the development of dichlorodiphenyltrichloroethane (DDT) and other organic

pesticides in 1947, arsenic was a common constituent of insecticides, herbicides,

fungicides and pesticides. These farming additives were the source of large-scale

arsenic contamination still persistent in the environment today. In the 1950’s, ‘gosio

gas’ was identified, a volatile, toxic arsenic species that originated from mould growing

on arsenic-containing pigments such as Scheele’s green (copper arsenite) used in

wallpaper colouring (Cullen and Reimer, 1989). Arsenic has long been used in the

preservation of wood products, generally in the form of CCA (copper, chromate and

arsenic) compounds. Loebenstein (1994) estimated that, during the early 1990’s, wood

preservatives were the dominant form of arsenic application in the industrial

environment. A decade later, the United States Environment Protection Authority

(USEPA), recognising the potential health implications of CCA treated products,

banned their use in the treatment of wood proposed for residential purposes.



Its description as a human carcinogen has prompted the recent reduction of arsenic use

in favour of less persistent and less toxic synthetic organic compounds. Nevertheless,

arsenic production has occurred for more than a century. The estimated amount of

arsenic derived from anthropogenic uses doubles that of naturally occurring arsenic

(ANZECC, 2000). The majority of anthropogenically produced arsenic comes as a by-

product from the smelting of metallic ores. Mines worked solely for arsenic are rare. In

fact, the Bolidan copper smelter in Sweden produces enough arsenic as a by-product to

supply the world’s entire demand (DMR, 1995). Han et al. (2003) provide a thorough

review of global industrial-age arsenic accumulation and production in the environment.

They calculated that in the year 2000, anthropogenic arsenic input into the world arable

surface was 2.18 mg arsenic kg-1, which is 1.2 times the abundance in the lithosphere.

Major anthropogenic arsenic sources were estimated to follow the order: As mining

production > As generated from coal > As generated from petroleum.

According to records held by the New South Wales Department of Mineral Resources

(DMR, 1995) arsenic production in Australia has been limited:

Prior to 1912 – small tonnages of ore were produced in South Australia and New

South Wales.

1916 to 1936 – major production occurred in several states for ‘prickly pear poison’

which was required in Queensland. All previous supplies had been shipped to

England during World War I. The New England region of NSW (encompassing

Stuarts Point) was an important mining area during this time.

1935 to 1949 – the market price for arsenic was low; the Wiluna mine in Western

Australia was the only Australian mine producing arsenic during this time.

Post 1952 – arsenic has not been produced in Australian mines since this time.

In NSW, arsenic ores have been recovered at eight main centres, with the New England

geological region accounting for 85% of the state’s production (DMR, 1995).

3.2 SOURCES OF ARSENIC IN THE ENVIRONMENT



3.2.1 Natural Occurrences Arsenic occurs rarely as the native metal, but is commonly associated with other

elements, particularly sulfide. Arsenic is a constituent of more than 245 minerals

(Bissen and Frimmel, 2003). As illustrated by the examples provided in Table 3.3, the

majority of arsenic minerals are associated with ore deposits and their alteration

products. Other trace elements, particularly other semi-metals and metals (such as Ag,

Mo, Pb, Sb, W) often occur in combination with these minerals.

Table 3.3 Common naturally occurring arsenic bearing minerals and their occurrence in the environment (adapted from Smedley and Kinniburgh, 2002; Palache et al., 1949).

Mineral Name Formula Occurrence

Adamite Zn2AsO4OH Oxidation zone mineral, forming near metal ore deposits

Arsenolite As2O3 Secondary mineral formed by oxidation of arsenopyrite, native arsenic and other As minerals

Arsenopyrite FeAsS Mineral veins

Claudetite As2O3 Secondary mineral formed by oxidation of realgar, arsenopyrite and other As minerals

Cobaltite CoAsS High-temperature deposits, metamorphic rocks

Glaucodote (Co,Fe)AsS Metalliferous veins with ores of tin, copper and silver – a replacement mineral of arsenopyrite

Loellingite FeAs2 Disseminated in veins near metal ores, difficult to distinguish from arsenopyrite

Native Arsenic As Hydrothermal veins

Niccolite NiAs Vein deposits and norites

Orpiment As2S3 Hydrothermal veins, hot springs, volcanic sublimation products

Pharmacosiderite

Fe3(AsO4)2(OH)3.5H2O

Oxidation product of arsenopyrite and other As minerals

Realgar AsS Vein deposits, often associated with orpiment, clays and limestones, also deposits from hot springs

Scorodite FeAsO4.2H2O Secondary mineral

The most abundant arsenic mineral is arsenopyrite (FeAsS) which contains 46% As. It

is generally assumed that arsenopyrite is formed under high temperature conditions

within the Earth’s crust, however, recent studies (Rittle et al., 1995) reported authigenic

arsenopyrite in sediments, while Newman et al. (1998) reported formation of orpiment



by microbial precipitation. Arsenic rich pyrite – Fe(S,As)2 – rather than arsenopyrite, is

likely to be the most important source of As in ore zones (Nordstrom, 2000). Arsenic

rich pyrite (called arsenian pyrite hereafter) is also suspected to be a constituent of acid

sulfate soil material.

In addition to arsenic mineral occurrences, volcanic eruptions; cycling from the oceans

(sea spray); wind erosion; low temperature volatilisation; and the burning of wood

through natural forest fires can also contribute arsenic naturally to the environment.

3.2.2 Anthropogenic Sources Section 3.1.3 discussed both former and current anthropogenic uses of arsenic.

Therefore, human production and use of arsenic and its compounds can lead to an

increased load of this element in the environment. Failure to adequately contain

arsenical compounds during or after their use may lead to mobilisation of arsenic into

environmental substrates. For example, the application of arsenic-containing

pesticides/herbicides to the pedosphere increases its mobilisation potential to underlying

groundwater, which may subsequently discharge to rivers or oceans, and become

available for bioaccumulation or methylation by marine organisms, eventually to be

recycled back to the pedosphere via precipitation. Its high mobility in the environment

ensures many substrates are affected by arsenic contamination.

Significant quantities of arsenic can be found in pesticides, insecticides, herbicides and

fertilizers. Mirlean et al., (2003) found that bioavailable arsenic in estuarine sediments

accounted for 80% of total arsenic and was derived from atmospheric contamination by

a fertilizer factory. Bioavailable forms can easily be transported into surface or

groundwater. Similarly, direct application of pesticides, insecticides and herbicides to

the land surface can enable infiltration of arsenic and other metals to groundwater,

unless plant uptake or surface run-off re-directs arsenic towards other substrates.

Mariner et al. (1996) investigated a groundwater plume derived from a

pesticide/herbicide chemical plant travelling towards the shores of Commencement Bay

in Washington. Chemical wastes from the plant had been placed in unlined pits,

generating an arsenic plume exceeding 250 mg L-1 As.



Arsenic is highly dispersed into the atmosphere through metal smelting processes and

the burning of fossil fuels. The burning of crude oil also contributes arsenic to the

atmosphere, to a lesser extent (Han et al., 2003). Fly ash from power plants may leach

arsenic into the surrounding soil and groundwater (Wallschlager et al., 2005). Mine

tailings can produce conditions suitable for arsenic mobilisation and transport in

aqueous environments. The most common arsenic associated ores are sulfides. Acid

mine drainage (AMD) results from extensive oxidation of Fe sulfide minerals and many

solid phase arsenic partitioning studies are conducted on the products of Fe sulfide

oxidation. One extreme case of AMD exists at Iron Mountain in the USA where pH

values are negative in AMD waters and arsenic concentrations reach the mg L-1 range

(Nordstrom et al., 2000).

Arsenopyrite and arsenian pyrite are often associated with ores of other metals, such as

Sn (Williams, 1997 as cited in Smedley and Kinniburgh, 2002); Au (Carrillo-Chavez et

al., 2000; Bodenan et al., 2004); Pb-Zn-Ag (Mok and Wai, 1990); Cu-W (Jung et al.,

2002); and Sb (Ashley et al., 2003; Mok and Wai, 1990). For example, arsenic derived

from stibnite deposits in the upper reaches of the Macleay River catchment has been

shown to contaminate both river water and stream sediments downgradient of the

Hillgrove stibnite mine (Ashley et al., 2003). In the Hetao Area of China, arsenic

concentrations in groundwater are also attibuted to downstream migration from mining

of metal sulfides and correspond to incidences of arsenic poisoning in local residents

(Zhang et al., 2002).

A less commonly documented occurrence of arsenic is its release after sand mining. The

same geochemical principles apply; arsenic is suggested to be released from the

oxidation of iron sulfides most likely present as pyrite in ASS. This suggestion was

made for the Tomago Sand aquifer (Coffey et al., 1996) located to the south of Stuarts

Point and was an instrumental finding leading to the inception of this project.

Other anthropogenic arsenic inputs include the disposal of industrial waste, leaching

from sheep dips and former tannery sites and contamination from CCA treated wood

products. CCA pressure treated wood has been widely used to preserve wood and

prevent decay by fungi, mould and termites. Children’s playgrounds, wooden decks and



utility poles are commonly constructed with CCA treated wood. Arsenic and Cu are

used as insecticides while Cr is added to fix As and Cu to the wood cellulose (Dawson

et al., 1991). Soil sampled close to CCA treated wood is normally high in As and

decreases with depth and spatial distance from the structure (see Chirenje et al., 2003

for an example). However, Chirenje et al. (2003) noted that this pattern only occurred in

new structures and suggested that soils surrounding older structures may have been

weathered and leached over time. This may increase the risk of arsenic entering local

groundwater resources.

Industrial anthropogenic activities may allow arsenic to enter groundwater via

atmospheric deposition, infiltration, streambed deposition and subsequent groundwater

recharge, or as direct contaminant sources to groundwater via spills or buried wastes.

Tanneries commonly use arsenic in the tanning process and may result in dissolved

arsenic plumes beneath the former/current tannery site. One such site has been

investigated by the author herein, situated in the Botany Sands aquifer of metropolitan

Sydney. Dissolved arsenic concentrations reach mg L-1 level and respond to rainfall

events in a pulsating manner (GHD, 2002). Davis et al. (1994) studied a highly reducing

groundwater environment downgradient of a historical tannery in Massachusetts.

Conditions were so reducing that DOC levels were extremely high (>100 mg L-1) from

hide breakdown, leading to reduction of As(V) to As(III) and subsequent methylation to

MMAA and DMAA. Sekhar et al. (2003) found arsenic in groundwater downgradient of

buried veterinary chemicals in India. Armienta et al. (1997) reported concentrations of

dissolved arsenic up to 1.097 mg L-1 in groundwater of the Zimapan Valley in Mexico.

Contribution via percolation of smelter fumes during the 1940’s was considered as a

potential arsenic source. In an extreme case of arsenic contamination, Jarmersted et al.

(1998) reported dissolved arsenic up to 100 mg L-1 downgradient of a sulfuric acid

production plant where arsenic and several other metals were present as by-products of

the sulfide ore roasting process.

3.2.3 Concentrations in the Environment Arsenic ranks 20th in abundance of elements within the Earth’s crust (NRC, 1977).

Table 3.4 indicates the variation in concentrations between natural versus anthropogenic

activities. A number of substrate examples have been provided in a series of different



settings to illustrate the potentially diverse range of arsenic concentrations that may be

encountered in the environment.

Table 3.4 Reported arsenic concentrations in various media.

MED

IUM

DESCRIPTION SAMPLE TYPE / LOCATION

CONCENTRATION (AVERAGE OR

RANGE)aREFERENCE

bulk continental crust general average 1.8 mg kg-1 Reimann and de Caritat (1998) lithosphere general average 1 Taylor and McLennan (1995) basalt general average 2 Drever (1997) Ocean ridge basalt general average 1 Reimann and de Caritat (1998) Gabbro, basalt general average 0.7 Reimann and de Caritat (1998) granite, granodiorite general average 3 Reimann and de Caritat (1998) granite general average 2 Drever (1997) granitic igneous general average 2 Langmuir (1997) shale, schist general average 13 Reimann and de Caritat (1998) shale general average 13 Drever (1997) shale/mudstone – marine general average 3 - 15 (up to 490) Smedley and Kinniburgh (2002) shale/mudstone – non-marine general average 3 - 12 Smedley and Kinniburgh (2002) shales and clays general average 10 Langmuir (1997) phyllite/slate general average 0.5 - 143 Boyle and Jonasson (1973) sandstone general average 0.5 Reimann and de Caritat (1998) sandstone general average 1 Drever (1997) sandstone general average 2 Langmuir (1997) limestone general average 1.7 Langmuir (1997) limestone general average 1 Drever (1997) limestone general average 1.5 Reimann and de Caritat (1998) coal general average 10 Reimann and de Caritat (1998)

RO

CK

coal general average 1,500 Bissen and Frimmel (2003) world soil mean 7.5 mg kg-1 Allard (1995) unpolluted soils Canadian range 4.8-13.6 Environment Canada (1993) surface soils US mean 7 Dragun (1991) uncontaminated soils - 0 - 150mm Australian average 7 Olszowy et al. (1995) uncontaminated soils - 0 - 100mm Australian average 3 Barry (1997) urban soil Trondheim 0.5 Reimann and de Caritat (1998) soil range Germany 2.5 - 15 Bissen and Frimmel (2003) soil range Italy 1.8 - 60 Bissen and Frimmel (2003) soil range USA 1.0 - 20 Bissen and Frimmel (2003) US Agricultural soils US mean <10 Kabata-Pendias and Pendias (1992)agricultural - surface Canada 6.6 Reimann and de Caritat (1998) agricultural - 0-25cm Finland 1.55 Reimann and de Caritat (1998) acid sulfate soils Vietnam 6 - 41 Gustafsson and Tin (1994) Acid sulfate soils Canada 1.5 - 45 Dudas (1984) deep sea clays average 13 Langmuir (1997) laterite Australia 3 Reimann and de Caritat (1998) polluted soils with smeltering Canadian range 50-100 Environment Canada (1993) soils on gold ore deposits Zimbabwe 20,000 Bissen and Frimmel (2003) soil near a lead smelter unknown 2,000 Bissen and Frimmel (2003) soil near a copper smelter unknown 550 Bissen and Frimmel (2003) soil near a gold smelter unknown 500 - 9,300 Bissen and Frimmel (2003) polluted soils with pesticides Canadian mean 54 Environment Canada (1993)

SOIL

polluted soils with wood preservatives

Canadian mean 6,000 Environment Canada (1993)

stream Austria 2 mg kg-1 Reimann and de Caritat (1998) stream Scotland 12 Reimann and de Caritat (1998) stream Germany 22 Reimann and de Caritat (1998) overbank Norway 5.3 Reimann and de Caritat (1998) river bed sediments Bangladesh 1.2 - 5.9 Datta and Subramanian (1997) SE

DIM

ENT

alluvial sand Bangladesh 1.0 - 6.2 BGS and DPHE (2001)



MED

IUM

DESCRIPTION SAMPLE TYPE / LOCATION

CONCENTRATION (AVERAGE OR

RANGE)aREFERENCE

alluvial mud/clay Bangladesh 2.7 - 14.7 BGS and DPHE (2001) US river and lake sediments mean 8.9 Kabata-Pendias and Pendias (1992)ocean sediments range 5-40 Neff (1997) sediments from gold ore deposits Zimbabwe 0.1 - 490 Bissen and Frimmel (2003) ocean world average 3 g L-1 Reimann and de Caritat (1998) ocean North Pacific 1 Reimann and de Caritat (1998) ocean - surface Atlantic 1.45 Carmalt and Norman (1998) ocean - surface Pacific 1.45 Carmalt and Norman (1998) ocean - deep Atlantic 1.53 Carmalt and Norman (1998) ocean - deep Pacific 1.75 Carmalt and Norman (1998) seawater average 0.09 - 24 (avg 1.5) Bissen and Frimmel (2003) Pacific Coastal Water Pacific Ocean 3 - 6 Ferguson and Gavis (1972) Northwest Pacific Pacific Ocean 1.2 Ferguson and Gavis (1972) Indian Ocean Indian Ocean 1.6 Ferguson and Gavis (1972) South West Indian Ocean Indian Ocean 3 Ferguson and Gavis (1972) seawater Coastal Australia 1.1 - 1.6 Maher (1985) seawater Coastal Spain 0.5 - 3.7 Navarro et al. (1993) ocean average 3 Drever (1997) English Channel England 2 - 4 Ferguson and Gavis (1972) freshwater average 0.15-0.45(max 1000) Bissen and Frimmel (2003) estuary Rhone, France 1.1 - 3.8 Seyler and Martin (1990) estuary Oslofjord, Norway 0.7 - 2.0 Abdullah et al. (1995) stream world average 4 Reimann and de Caritat (1998) stream Nova Scotia 0 Reimann and de Caritat (1998) streams average 2 Drever (1997) river Sweden 0.2 - 0.4 Ferguson and Gavis (1972) river Japan 0.25 - 7.7 Ferguson and Gavis (1972) Elbe River Germany 20 - 25 Ferguson and Gavis (1972) Columbia River USA 1.6 Ferguson and Gavis (1972) lake Greece 1.1 - 54.5 Ferguson and Gavis (1972) lake Japan 0.16 - 1.9 Ferguson and Gavis (1972) rain - coastal Norway 0 Reimann and de Caritat (1998) rain - inland Norway 0 Reimann and de Caritat (1998) rain on land unknown 1.6 Wedepohl (1969) rain at sea unknown 0.6 Wedepohl (1969) groundwater Norway 0 Reimann and de Caritat (1998) surface/ground water median 2 Langmuir (1997) groundwater average UK 0.5 - 10 Edmunds et al. (1989) As-rich groundwater Bengal Basin,

Argentina, Mexico, China, Taiwan and

Hungary

10 - 5,000 Das et al. (1995), BGS and DPHE (2001), Nicolli et al. (1989), Smedley et al. (2001), Del Razo et al. (1990), Luo et al. (1997), Hsu et al. (1997), Varsanyi et al. (1991)

acid mine drainage 7 countries from SE

Asia, Africa, Latin America

5 - 72,000 Williams (2001)

WA

TER

geothermal waters Old Faithful, Yellowstone

National Park

1,275 Stauffer and Thompson (1984)

world - remote average 0.5 ng/m3 Reimann and de Caritat (1998) average unpolluted estimate average 0.1 - 10 WHO (2001) average urban estimate average 3 - 180 WHO (2001) polluted area estimate > 1,000 WHO (2001) A

IR

world - polluted average 15 Reimann and de Caritat (1998) a Concentrations reported in mg kg-1 (rocks, soil and sediment), g L-1 (water) and ng/m3

(air).



3.2.3.1 RocksArsenic concentrations in igneous rocks are generally low (less than 2 mg kg-1 As).

Metamorphic rocks tend to reflect concentrations present in their igneous and

sedimentary predecessors, with average concentrations from Table 3.4 also indicating

less than 2 mg kg-1 As. Pelitic rocks derived from metamorphosed and non-

metamorphosed argillaceous rocks (shales, phyllites, slates) report slightly higher

averages of around 13 mg kg-1 As, but can reach much higher levels depending on their

depositional environment. Concentrations in sedimentary rocks vary due to major

constituents, sediment provenance and particle size. Sandstones, dominated by quartz

and feldspars, contain lower average arsenic concentrations (1-2 mg kg-1 As) than shales

(13 mg kg-1 As) which exhibit more clay, organic matter, oxide and sulfide minerals.

Coal can reach concentrations higher than 1,500 mg kg-1 As depending on the amount

of impurities present.

3.2.3.2 SoilsConcentrations of arsenic in soil will vary depending on parent material, grain size, soil

constituents, proximity to mineralised zones and anthropogenic influence. Values

shown in Table 3.4 range from 0.5 – 60 mg kg-1 As for suspected unpolluted soils; less

than 10 mg kg-1 As for agricultural soils; up to 45 mg kg-1 As in acid sulfate soils; and

excessively high concentrations in soils suspected of being anthropogenically

contaminated.

3.2.3.3 SedimentsSediment constituents play a vital role in determining arsenic concentrations. There is

no significant difference between sediments and rock concentrations containing the

same major constituents. Clays, muds and sediments high in oxides and sulfides will

retain more arsenic than sands and carbonates. Increased concentrations may occur due

to mineralised zones and anthropogenic sources.

3.2.3.4 WaterOcean waters typically average ca. 3 g L-1 As in the dominant form of As(V). This

figure applies to estuarine waters also, but may vary slightly due to salinity and redox



gradients. Fresh water in rivers and streams are slightly lower (around 1 g L-1 As) but

are highly influenced by drainage areas, baseflow and point contamination sources such

as processing plants located on riverbanks. Rivers draining As-poor bedrock will have

lower As concentrations in their waters (less than 1 g L-1 As) compared to rivers

draining As-rich bedrock or geothermal areas. Nimick et al. (1998) found up to 370 g

L-1 As in the Madison River as a result of geothermal inputs from the Yellowstone

geothermal system. Lake water As concentrations can be influenced by stratification,

evaporation and seasonal effects. Changes in redox gradients with depth can lead to

stratification of lakes and influence on arsenic species available for sorption onto iron

oxides. Azcue and Nriagu (1995) studied arsenic distribution in lake waters from

Ontario and found that seasonal biological activity in the lake accounted for differences

in arsenic species distribution in summer months. Closed basins can account for high

evaporation rates and elevated arsenic levels in some lakes. Rain water arsenic

concentrations are typically below 1.5 g L-1 but may differ slightly between oceanic

and terrestrial precipitation localities.

Background As concentrations in groundwater are greatly influenced by the

composition of the aquifer matrix. Natural concentrations are reported in the range 0 –

5,000 g L-1 in the literature, but typically occur as less than 10 g L-1 as provided by

Edmunds et al. (1989) for the UK, and Welch et al. (2000) for the USA. Arsenic-rich

groundwaters are widely dispersed throughout the world and often occur as a result of

natural processes. Mining, agricultural and industrially influenced arsenic groundwaters

can exhibit very high arsenic concentrations (in the mg L-1 range) but tend to be more

localised than the naturally occurring systems.

3.2.3.5 AirAtmospheric arsenic concentrations are typically in the order of a few nanograms per

cubic metre. Elevated levels are largely influenced by proximity to industrial emitting

infrastructure such as smelters and fossil fuel power plants, or volcanic activity. Much

of the atmospheric arsenic is particulate and is cycled back to the lithosphere through

wet and dry deposition.



3.3 AQUEOUS ARSENIC CHEMISTRY

Arsenic primarily exists as As(III) or As(V) in dissolved form. Its speciation is largely

controlled by changes in redox potential of the surrounding solution. In oxidising

conditions (generally >+100mV) As(V) dominates as deprotonated oxyanions of arsenic

acid (Figure 3.1). In reducing conditions (<+100mV) arsenious acid (As[III]) exists in

its various deprotonated forms. At Eh values below –250mV insoluble arsenic

compounds such as As2S3 can form when in the presence of sulfur or hydrogen sulfide.

These compounds are insoluble under neutral or acidic conditions. Arsine gas and free

elemental arsenic are formed only under extremely reducing conditions (Bissen and

Frimmel, 2003).

The deprotonation of arsenic species is similar to carbonic acid; it releases protons

stepwise as the pH of the solution increases. For arsenite, the fully protonated form is

H3AsO30 and the first dissociation (pK1) occurs at pH 9.2 (equation 3.1);

H3AsO30 H2AsO3

- + H+ (3.1)

Likewise, for arsenate, the fully protonated form is H3AsO40 and the first dissociation

(pK1) occurs at pH 2.2 (equation 3.2);

H3AsO40 H2AsO4

- + H+ (3.2)

Table 3.5 lists the remaining dissociation constants while Figure 3.2 illustrates these

reactions diagrammatically.

Table 3.5 Dissociation constants for protonated arsenite and arsenate. Species pK1 pK2 pK3

H3AsO30 [As(III)] 9.2 12.1 13.4

H3AsO40 [As(V)] 2.2 7.0 11.5




Figure 3.1 Eh-pH diagram for the As-O2-S-H2O system at 25ºC and 1 bar pressure (Bissen and Frimmel, 2003).

Arsenic (III) is a neutral, uncharged molecule (H3AsO30, pKa = 9.2) at the pH of most

natural waters, making it easily mobile as this neutral charge is less likely to sorb

strongly to mineral surfaces (Korte and Fernando, 1991). This may explain its increased

mobility in comparison to As(V) which sorbs more strongly under the pH of most

natural waters. Arsenic exhibits a relatively slow redox transformation (Masscheleyn et

al., 1991) and is often therefore responsible for the presence of both As(III) and As (V)

in the same redox environment. Smedley and Kinniburgh (2002) explain that

equilibrium thermodynamic calculations predict As(V) should prevail over As(III) in all

but strongly reducing environments where sulfate reduction is occurring. However,

catalysis of arsenic redox reactions often occurs by interaction with micro-organisms.

Bacterially mediated arsenic redox reactions can be orders of magnitude quicker than

chemical oxidation and reduction (Smedley and Kinniburgh, 2002) and are thus gaining

newfound importance in many groundwater remediation investigations. Cullen and

Reimer (1989) provide a review of the interaction of micro-organisms with arsenic



compounds, while Ahmann et al. (1997) and Oremland and Stolz (2005) investigate

these interactions in groundwater environments.


Figure 3.2 Deprotonation of (a) As(III) and (b) As(V) species over different pH ranges (Smedley and Kinniburgh, 2002).

Smedley and Kinniburgh (2002) provide a good review of the distribution of arsenic

species in various water bodies. As(V) generally dominates in surface waters but can be

affected by stratification, anoxic bottom sediments, microbial interaction and

anthropogenic input. In relation to groundwaters, they note the ratio of As(III) to As(V)

can vary greatly as a result of redox conditions, microbes and diffusion. In the reducing

groundwaters of Bangladesh As(III)/Total As ratios are typically between 0.5-0.6.



3.4 IMMOBILISATION OF ARSENIC: SOLID PHASE RETARDATION

Arsenic (and other trace elements) associated with the solid phase may be present in a

variety of different chemical forms – precipitated as a mineral phase; adsorbed onto a

mineral surface; incorporated within the crystal structure of a mineral or partitioned into

organic matter. A general term applied to all of these processes is sorption, and a brief

review of these mechanisms is provided below. A more specific discussion on

documented arsenic sorption phenomena follows which examines experiments proven

to constrain arsenic mobility and retain it in the solid phase.

3.4.1 Properties of Sorbent Materials

3.4.1.1 Surface Area Solid materials in aquifers can act as sorbents, with strength of sorption dependent on

both the mineralogy and the surface area of the solid. Larger surface areas sorb more,

making smaller grain size fractions (such as clays) responsible for much of the sorption

in aquifers. However, oxide or organic coatings on the coarser (i.e., larger) sand grains

can also sorb material. Therefore, the content of clay, type of clay, amount of organic

matter present and surface coatings can all contribute to sorption capacity in sediments.

3.4.1.2 Surface charge Concurrently, an increase in surface area of a mineral also increases its surface charge

(number of charged sites per unit area, commonly referred to as surface site density)

which are both directly proportional to the sorptive properties of a particular mineral

(Table 3.6). Clay minerals exhibit high surface areas in comparison to oxides of iron,

manganese and aluminium; however, oxides are characterised by high surface site

densities increasing the likelihood of sorption.

A mineral surface can exhibit two types of surface charge – permanent or variable.

Permanent surface charge arises from the isomorphic substitution of lower valence

cations in the crystal lattice of minerals. For example, Al3+ replaces Si4+ in the

tetrahedral layer and Mg2+ replaces Al3+ in the octahedral layer of clay minerals and

zeolites resulting in a net negative charge. An excess of O2- bonds creates this negative



surplus. Similarly, lattice imperfections in mineral structures, such as a deficit of Al3+ in

the octahedral layer or lack of interlayer K+, can lead to a net negative surface charge

(Langmuir, 1997). Isomorphic substitutions and lattice imperfections are permanent

surface charges created when the mineral was formed and are independent of solution

chemistry.

Table 3.6 Measured or estimated surface areas and surface site densities for different geological materials (table adapted from Langmuir, 1997).

Mineral / Phase Surface Area (m2 g-1) Surface Site Density (sites per area)

quartz 0.14 4.2 to 11.4/nm2

rutile 5 to 19.8 5.8/nm2

hematite 1.8 5 to 22/nm2

goethite 45 to 169 2.6 to 16.8/nm2

ferrihydrite 250 20/nm2

birnessite 290 18/nm2

gibbsite 120 2 to 12/nm2

kaolinite 10 to 38 1.3 to 3.4/nm2

illite 65 to 100 0.4 to 5.6/nm2

montmorillonite 600 to 800 (especially interlayer)

0.4 to 1.6/nm2

Organic substances such as humic materials

260 to 1,300 2.31/nm2

Variable surface charge is caused by broken or unbalanced bonds on the surface of the

mineral leading to ionization of surface functional groups such as O2- and OH- and

generally resulting in a net negative surface charge. These charges are pH-controlled

due to the increased H+ adsorption onto negative surface sites as pH decreases. Surface

functional groups can be organic or inorganic molecular units (Sparks, 2003). The pH at

which all surface functional groups are neutralised is called the point of zero charge

(PZC) and is specific for individual minerals.

Most clays exhibit constant negative charge due to isomorphic substitution or lattice

imperfections. This is particularly important for the illite, smectite and vermiculite clay

minerals. Broken bonds at clay surfaces can also cause variable negative charges – this

is the chief form of surface charge for kaolinite, but much less for other clay minerals

which generally exhibit permanent negative charges. Variable surface charge is an

important function of oxyhydroxides, kaolinite, phosphates and carbonates.



3.4.2 Surface Sorption The term sorption applies to both 2D (adsorption onto a mineral surface) and 3D

(absorption into the mineral structure) processes. It is frequently used in geochemical

literature because the actual type of sorption process may not be known. Figure 3.3

illustrates the various types of sorption mechanisms possible at the mineral/water

interface, some of which are further explained below.


Figure 3.3 Mechanisms of sorption at the mineral/water interface. (a) adsorption of an ion via formation of an outer-sphere complex; (b) loss of hydration water and formation of an inner-sphere complex; (c) lattice diffusion and isomorphic substitution within the mineral lattice; (d) rapid lateral diffusion and formation either of a surface polymer or (e) adsorption on a ledge which maximizes the number of bonds to the atom; (f) upon particle growth surface polymers end up embedded in the lattice structure; and (g) the adsorbed ion can diffuse back into solution either as a result of dynamic equilibrium or as a product of surface reactions (from Charlet and Manceau, 1993; as cited in Sparks, 2003).

3.4.2.1 Surface Complexation Surface functional groups will competitively adsorb metal ions, protons, and oxyanions

when pH conditions provide suitable surface charges. For example, surface hydroxyl

groups (-OH) present on an iron oxide surface ( Fe) may complex with cations in the

following manner (equation 3.3):

Fe-OH- + Cu2+ Fe-OCu+ + H+ (3.3)



or, in a similar manner, anions can complex (equation 3.4):

FeOH+ + H2AsO4- FeHAsO4

- + H2O (3.4)

Similar adsorption mechanisms have been proposed for anions on sulfide minerals. The

overall reaction is called a ligand-exchage reaction or surface complexation. Two types

of surface complexation exist; outer-sphere and inner-sphere. According to Sposito

(1989) if a water molecule is present between the surface functional group and the

bound ion or molecule, then it is an outer-sphere complex (sometimes called an ion

pair). The bonds between the ligand and the central metal ion are generally electrostatic

in nature. This type of reaction is particularly important in waters with high ionic

strengths, such as seawater (Morel and Hering, 1993). If no water molecule is present, it

is an inner-sphere complex (Figure 3.4).


Figure 3.4 (a) Inner-sphere complex formation and (b) Outer-sphere complex formation (taken from Sposito, 1984; as cited in Sparks, 2003).

For outer-sphere complexes, anion sorption is favoured when the solution pH is below

the PZC of the mineral surface and cation sorption is favoured when the pH is above the

PZC. For example, the PZC for hydrous ferric oxides is 7.9-8.2 (Dzombak and Morel,

1990) making arsenic adsorption important in acidic and near-neutral solutions. Thus

equation (3.4) above would be important at pH’s below 8.2 if arsenate is sorbed as an

outer sphere complex. To complicate the matter, arsenic sorption via inner-sphere

complexation on hydrous ferric oxides produces stronger binding associations. This

enables As(V) to adsorb well above the PZC of hydrous ferric oxides (Romero et al.,

2004). Predicting which type of surface complexation may occur involves a good



knowledge of the experimental literature, characterisation of the matrix and various

potential geochemical interactions that may influence sorption at the mineral interface.

For a broad sense of sorption occurrence, Romero et al. (2004) suggest that inner-sphere

complexation appears to take place at high concentrations of dissolved metals, while

outer-sphere complexation (electrostatic adsorption) generally dominates at low metal

concentrations.

3.4.2.2 Permanent versus Variable Surface ChargeThe ion exchange capacity is a measure of a mineral to sorb cations/anions to their

surface and can incorporate both permanent (isomorphous) and variable charges. In

assessing the Cation Exchange Capacity (CEC) of a material, solution pH carries more

dependence to those materials where surface functional groups exist and variable charge

is pH controlled (Table 3.7). Thus the CEC for kaolinite increases with increasing

surface area of a mineral (i.e. decreasing particle size exposes more surface area and

hence more surface functional groups). However, there is no marked difference for the

other clay minerals which are predominantly controlled by permanent (isomorphic

substitution) surface charges (Sparks, 2003) and therefore are less influenced by

changes in mineral surface area. For anions, oxide surfaces and organic matter tend to

exhibit higher affinities for arsenic than most clay minerals.

Table 3.7 CEC’s of different geologic materials measured at pH 7. Absent, negligible or slight pH dependencies indicate the material obtains its CEC primarily via isomorphous substitution. Strong dependencies indicate the surface charge results from adsorption at the mineral surface (table adapted from Langmuir, 1997).

Material CEC (meq/100g) pH Dependence Kaolinite 3 to 15 Strong

Glauconite 11 to 20 Slight Illite and Chlorite 10 to 40 Slight

Smectite-montmorillonite 80 to 150 Absent or negligible Vermiculite 100 to 150 Negligible

Zeolites 100 to 400 Negligible Organics in soils, humic matter 100 to 500 Strong

Mn(IV) and Fe(III) oxyhydroxides 100 to 740 Strong



So, when the pH of the surrounding solution is high the surface carries a net negative

charge and is conducive to cation exchange. When the pH of the surrounding solution is

low the surface carries a net positive charge providing conditions suitable for anion

exchange. This process may play an important part in this study given that arsenic forms

negative oxyanions in solution which will be attracted to net positive surface charges.

Clays with high CEC’s can remove various cations from solution and retain them in the

solid phase.

3.4.3 Arsenic Retention in the Solid Phase Having reviewed the processes responsible for ion sorption we now focus on the

behaviour of potential sorbents available for arsenic retardation at Stuarts Point. The

heterogeneity of the aquifer matrix is conducive to many different arsenic

immobilisation processes, including mineral precipitation as well as sorption; with both

processes influenced by the surrounding water chemistry. The dominant arsenic-specific

immobilisation processes are presented here.

3.4.3.1 Solid Phase Arsenic FormationArsenic bearing minerals have previously been discussed and their distribution is a

product of well documented ore genesis processes. Geological mapping and

geochemical tracing techniques can identify zones of arsenic minerals and its elemental

occurrence in association with various other mineral deposits. Geochemical conditions

must be suitable for ore genesis, with shallow unconsolidated sandy aquifers (such as

Stuarts Point) rarely providing these conditions for arsenic mineral precipitation.

Precipitation of arsenic minerals occurs when geochemical conditions are suitable and

sufficient concentrations of respective elements are present in dissolved form.

Numerous theoretical and experimental studies exist on the precipitation of arsenic with

various elements, notably sulfides and metal arsenates (Wagemann, 1978; Crecelius et

al., 1986) including calcium and magnesium (Raposo et al., 2004). Saturation of the

aqueous solution with respect to mineral solubility will promote precipitation.

Geochemical modelling codes generally contain thermodynamic databases that are

incorporated into many studies and designed to assess the plausibility of arsenic mineral

precipitation. These databases should be thoroughly reviewed before use so as to



include the most recent thermodynamic data applicable to the system being modeled.

Failure to do this can result in unrealistic results, as was the case adopted by Crecelius

et al. (1986) who found barium arsenate was oversaturated in the solution being

modeled. They attributed this to an error in the thermodynamic data as there was no

geological occurrence of the material and dissolved barium would have been

analytically undetectable. An investigation by Planer-Friedrich et al. (2001) using an

updated Saturation Index (SI) for BaHAsO4.H2O (rather than Ba2[AsO4]2) produced

results which suggested barium arsenate may precipitate but only at higher barium

concentrations (up to 170 g L-1) than those used in their study. They thus attributed the

negative arsenic and barium correlation to differences in geological settings. To solve

the thermodynamic question regarding barium arsenate stability in natural waters, Zhu

et al. (2005) conducted specific experimental investigations to conclude that barium

arsenate was unlikely to form under natural conditions in water. Their updated

thermodynamic data should be included in all geochemical databases so as to avoid

misconceived results.

Perhaps a mechanism of more recently examined arsenic mineral association in

groundwater resource investigations is that of its association with pyrite. Pyrite is

formed when water containing iron and sulfate, in the presence of organic matter,

crystallises to form secondary iron sulfides (Fitzpatrick et al., 1996). Sulfate reducing

bacteria, such as Desulfovibrio desulfuricans, use sulfur as an electron sink during the

oxidation of organic matter, reducing sulfate to sulfide which then combines with

ferrous iron under anaerobic environments (Fitzpatrick et al., 1996). This process can

occur in both marine and non-marine environments. Coastal pyrite is often associated

with acid sulfate soil formation. Authigenic pyrite is common where marine incursion

has occurred in the sediments, such as marine shales (Craw and Chappell, 1999). Recent

studies in Australia (Fitzpatrick et al., 1996) have identified pyrite formation under

inland dryland salinisation processes, whereby saline sulfate groundwater combines

with iron and organic matter to produce suitable pyritization conditions. Pyritization in

salt marsh sediments can be rapid; in other sedimentary environments the process is

relatively slow as amorphous iron monosulfides (FeS) react with elemental sulfur (S0) to

form pyrite (Berner, 1970).



Early research on the association of arsenic with pyrite includes a study of marine pyrite

capture of arsenic in the Laurentian Trough of Canada (Belzile and Lebel, 1986). Iron

extracted from the targeted pyrite fraction during sediment extraction procedures

produced a correlation of 0.981 with arsenic also extracted from this phase. This early

work suggested that arsenic may be present within the lattice sites of pyrite. Research

on acid sulfate soil by-products generated further interest in the association of arsenic

with pyrite. In 1984 and 1987 Dudas studied the relationship between arsenic and pyrite

in acid sulfate soils. He found that although pyrite was already known to be a carrier of

arsenic, containing as much as 5,600 mg kg-1 As (Boyle and Jonasson, 1973), there was

no information on the arsenic chemistry of ASS. Gustaffson and Tin (1994) confirmed

Dudas’ earlier correlative work and also suggested that arsenic is predominantly

redistributed in iron oxide phases after oxidation of arsenic-rich pyrite. Still no

definitive studies had been produced on the detailed mechanisms of arsenic formation

with pyrite.

Belzile (1988) examined the fate of arsenic in estuarine and coastal offshore sediments,

finding that the formation of pyrite greatly affected the distribution of arsenic. He once

again proposed that arsenic was included into the lattice sites of pyrite, in a constant

Fe/As ratio of 1000. Previous Fe/As ratios were established by Boyle and Jonasson

(1973) to range from 290 to 4,660 in rocks containing pyrite. Belzile’s (1988) constant

Fe/As ratio could reflect the maximum arsenic retention capacity for the framboidal

pyrites he examined, but may not be reproduced in the same ratio in other environments.

Much of the work following these earlier studies focussed on contaminated sediments

and the association of arsenic with various selectively extracted soil fractions, with the

pyrite fraction often specifically examined (Gruebel et al., 1988; Lacal et al., 2003; La

Force et al., 2000; Moore et al., 1988; Van Herreweghe et al., 2003; Wenzel et al.,

2001). While this work has been important in deciphering the correlation between

arsenic and iron sulfide phases it has not been successful in delineating specific sorption

mechanisms between arsenic and pyrite.

By the 1990’s it became commonly accepted that arsenic is isomorphically substituted

(sometimes called solid-solution) for reduced sulfur in the pyrite structure (Presser and



Swain, 1990; Nesbitt et al., 1995; Strawn et al., 2002). The most recent studies are now

examining these mechanisms via the use of spectroscopic methods in order to determine

arsenic local structure and oxidation state. Savage et al. (2000) used electron

microprobe, transmission electron microscope and extended X-ray absorption fine-

structure spectroscopy (EXAFS) to indicate that As substitutes for S in pyrite,

producing a mineral commonly referred to as arsenian pyrite. Their detailed studies of

mine tailings on a molecular level indicated that the amount and spatial arrangement of

arsenic substitution for sulfur can lead to enhanced weathering of arsenian pyrite in

comparison to pure pyrite phases. Zacharias et al. (2004) concurrently confirmed this

most recent work by providing electron microscope quantitative analyses of pyrites

from a granitoid intrusion in Prague. The atomic percentages of Fe, S and As indicated

an arsenic association with pyrite via sulfur substitution – Fe(AsxS1-x).

The common relationship between the presence of dissolved arsenic and sulfide

minerals in anoxic environments led Bostick and Fendorf (2003) to examine in detail

the mechanisms by which sulfide minerals sequestered arsenic under such conditions.

At pH levels above 2-3 iron sulfide minerals are negatively charged (Bebie et al., 1998).

In contrast to arsenic adsorption onto iron oxides (which is low at high pH levels),

arsenite sorption on pyrite and troilite (FeS) increased at higher pH. The authors

proposed that strong, inner-sphere complexes sorbed the arsenite to the iron sulfides,

rather than outer sphere complexation mechanisms. In contrast to isomorphic

substitution, they also suggested the formation of an intermediate surface precipitate

with a structure resembling arsenopyrite (FeAsS). Initially, arsenite was reduced to an

FeAsS-like precipitate on both FeS and FeS2. The corresponding oxidation of Fe(II) and

sulfide was detected but low amounts were observed in dissolved phase. Bostick and

Fendorf (2003) subsequently proposed the following reaction (equation 3.5) for arsenic

surface precipitation on iron sulfides:

3FeS + As(OH)3 FeS2 + FeAsS + Fe(OH)3 (3.5)

This reaction supports the association of arsenic and pyrite observed via the numerous

selective extraction techniques found in the scientific literature. In both sorption and

surface precipitate cases an insoluble arsenic-bound sulfide phase is formed that is



effective in immobilising arsenic in reducing sulfidic environments. Oxidation of these

phases are predicted to release arsenic species from their sorbed sulfides, similar to the

reductive dissolution of iron oxide phases releasing arsenic into dissolved phase.

Follow-up studies conducted by Bostick et al. (2004) confirmed that sorption reactions

on sulfide minerals in anoxic environments are the initial step in arsenic sequestration

processes. Precipitation of an FeAsS phase on sulfide minerals, specifically pyrite,

eventually led to conversion of a more stable arsenic sulfide mineral (As2S3) in sulfidic,

reducing environments.

3.4.3.2 Surface Reactions Affecting Arsenic Mobility Arsenic sorption to solid phases is a complex mechanism currently receiving an

increased amount of research attention. Experimental studies using pure mineral phases

and high powered microscopic techniques are providing valuable information for

geochemists, material scientists, environmental scientists and pedologists alike. Using

the data obtained through these theoretical investigations, predictions regarding arsenic

retardation in heterogenous matrices can be made.

Arsenic sorption in such highly composite settings can be controlled by mineral type,

crystallinity, surface features and properties of the surrounding solution. A number of

arsenic mineral sorption studies are reviewed in order to show the wide range of

potential conditions suitable for arsenic surface sorption in the environment.

3.4.3.2.1 SURFACE COMPLEXATION (ADSORPTION)

Lin and Puls (2003) found the following order of decreasing arsenic affinity on various

minerals: iron hydroxides > clays > feldspars. This illustrates the complexity of arsenic

adsorption onto various mineral substrates which are discussed independently below.



Oxyhydroxides1

Common oxides of Fe, Al, Mn and silica are widely dispersed in nature as fine particles

and surface coatings. They have an excellent ability to sorb cations and anions to their

surface thereby decreasing the mobility of many trace elements in the environment.

Given the high affinity of Fe-oxides for arsenic adsorption much work has focussed on

them, however similar results have also been reported for other oxides.

Both crystalline and amorphous iron oxide forms have hydroxylated surfaces allowing

surface adsorption processes to occur (Dudas, 1987). The amount of arsenic sorbed to

an iron oxide surface, however, is dependent on both arsenic speciation and mineral

crystallinity. Iron oxides initially form poorly crystalline (amorphous2) phases with high

surface areas which undergo transformation over time to more crystalline phases such as

goethite, hematite or magnetite (Dixit and Hering, 2003). With the advent of

crystallisation comes a loss in specific surface area, however, Dixit and Hering (2003)

found that the transformation of HFO to goethite did not decrease the affinity of the

sorbed phase for arsenic. Nevertheless, reduction in surface size and site density during

crystallisation may increase the mobilisation of any surface-bound arsenic.

Arsenic adsorption is more highly dependent on its oxidation state than on pH in the

range of 5.5-7.5 (Raven et al., 1998). Several studies have been conducted on the

adsorption of arsenic to ferrihydrite (Raven et al., 1998; Jain et al., 1999; Appelo et al.,

2002); goethite (Grafe et al., 2001); and hematite (Redman et al., 2002). Results of these

investigations vary according to experimental conditions. The general consensus for

optimum arsenic adsorption onto ferrihydrite and goethite under different pH conditions

is summarised in Table 3.8.

1 The term ‘oxyhydroxide’ will be used herein to collectively describe the terms hydroxide, oxide and oxyhydroxide, unless specifically stated. 2 Iron (hydr) oxides, hydrous ferric oxides (HFO) or hydroxylated iron oxides will generally be termed as amorphous or specified as ferrihydrite herein.



Table 3.8 pH concentrations required for optimum arsenic sorption conditions onto different iron oxides.

pH Conditions Required for Optimum Arsenic Sorption Iron Oxide As (III) As (V)

Ferrihydrite a 6-9 4 Goethite b 6-9 3

a Dixit and Hering (2003) b Grafe et al. (2001)

According to the literature, the maximum adsorption of As(III) generally corresponds

with the first pKa (9.2) of H3AsO30 while for As(V) maximum adsorption normally

occurs close to its first pKa (2.2) at a pH of 3-4 (Raven et al., 1998; Grafe et al., 2001).

The lower adsorption of As(V) at high pH levels can be attributed to an increased

amount of electrostatic repulsion between increasingly negative arsenate oxidation

states and the negatively charged iron oxide surface. Arsenite species present at the

same pH levels as the arsenate species are less negatively charged thereby not

exhibiting as much increased repulsion. Hence adsorption of As(III) decreases less with

increasing pH (Raven et al., 1998) when compared to As(V).

Dzombak and Morel (1990) provide a database for sorption onto HFO by various ions.

Data available for anion sorption was sparse compared to cation sorption, however is

sufficient to provide an indication of basic trends in anion sorption behaviour. Dzombak

and Morel (1990) used three data sets (Davis and Leckie, 1980; Pierce and Moore, 1980

and 1982) to indicate arsenic adsorption onto HFO. Proposed surface complexation

reactions for each species are provided in Table 3.9 while Figure 3.5 suggests arsenate

and arsenite adsorption behaviour over different pH levels that appear to be consistent

with results reported from the literature.

Table 3.9 Surface complexation reactions for arsenate and arsenite on HFO (data derived from Dzombak and Morel, 1990).

Arsenate FeOH0 + AsO4

0 + 3H+ = FeH2AsO40 + H2O K1

FeOH0 + AsO4- + 2H+ = FeHAsO2

- + H2O K2 FeOH0 + AsO4

2- + H+ = FeAsO42- + H2O K3

FeOH0 + AsO43- = FeOHAsO4

3- K4Arsenite

FeOH0 + H3AsO30 = FeH2AsO3

0 + H2O K1



By reference to the above cited literature, inner-sphere complexation has been proposed

for the dominant arsenic sorption mechanism onto metal oxides via displacement of a

hydroxyl group on the oxide surface. Sherman and Randall (2003) agreed that evidence

for inner-sphere complexation of arsenic is sufficient but that the nature of the surface

complexes remains controversial. Consistent with sorption mechanisms for iron oxides,

both aluminium and manganese oxides produce inner-sphere surface complexation

reactions with arsenic thereby reducing its mobility under suitable aqueous geochemical

conditions.


Figure 3.5 Arsenic adsorption on HFO for a) As(V) and b) As(III) according to ionic strength (I), Fe and As concentrations given (after Dzombak and Morel, 1990). Solid line is optimal fit, dashed line is an estimate using sorption constants.

Clay Minerals

As noted previously, the clay mineral type, surface charge and surface area, in addition

to the presence of any other trace elements, may all influence arsenic sorption onto clay

mineral surfaces. While many studies have been conducted on arsenic affinity for iron

oxide adsorption, clay mineral adsorption has been studied significantly less despite

their abundance in aquifer sediments, soils and particulate matter. Two predominant

experimental studies on arsenic sorption with clay minerals are frequently cited in

arsenic geochemical reviews.

Lin and Puls (2000) investigated adsorption, desorption and oxidation of arsenic

affected by clay minerals and aging. Six different clay minerals were utilised in their

study; halloysite, two different kaolinites, illite, an illite/montmorillonite mix and



chlorite. Arsenic (V) exhibited higher adsorption affinity than As(III) for all the clay

minerals studied. This adsorption affinity was influenced by pH; As(V) adsorption

decreased at pH 7.5 (compared to experimental pH’s of 5.5 and 6.5) while As(III)

increased at pH 7.5. This difference can be attributed to the pKa values of arsenic acid

(H3AsO4 – As[V]) ranging from 2.2 to 11.5 in comparison to arsenous acid (H3AsO3 –

As[III]) ranging from 9.2 to 13.4. Therefore, at pH 5.5 to 7.5 the positively charged

particle edges would be more conducive to the lower pKa values assigned to the As(V)

protonated species (2.2 – 11.5) over the As(III) species.

Desorption of both As(III) and As(V) decreased over time, indicating aging could

hinder the desorption of arsenic. Two theories were presented to explain this

phenomena:

1. Arsenic adsorbed onto external clay surfaces may diffuse into internal pores; and

2. Loss of hydration water.

Both theories contribute to the strengthened bonding of arsenic with the clay mineral

and promoting less desorption of arsenic. Aging over time also allowed oxidation of

As(III) to take place resulting in significant amounts of surface-bound As(V). Oxidation

effects are important due to the increased affinity of As(V) for surface sites over the

more mobile and toxic As(III) species. The mechanism of oxidation is not well

understood but may be due to the presence of other elements within the clay mineral

structure.

Individually, each clay mineral exhibited different sorption characteristics. The illites

and montmorillonites are 2:1 layer clays whose negative surface charge is largely

controlled by isomorphic substitution in the lattice. Positive surface charges for anion

adsorption can be generated by protonation on broken Al-OH bonds exposed at the

particle surface. Chlorite exhibits similar features. In contrast, the kaolinites and

halloysite are 1:1 layer clays and do not demonstrate extensive isomorphic substitution

– broken edge effects are thus more important for these minerals.

The illite and mixed illite/montmorillonite in Lin and Puls’ (2000) study had moderate

arsenic adsorption and subsequently moderate desorption. The exposed Al-OH edge

groups were postulated to control these results and hence influence arsenic sorption with



changes in pH (because variable surface charges are pH controlled). Kaolinite,

theoretically predicted to sorb and desorb more arsenic than these clays, produced

similar sorption/desorption results. The authors attributed this anomaly to the fine-

grained nature of the illite and montmorillonite samples used in the experiment which

may have been prone to increased surface reactivity.

Halloysite displayed the greatest adsorption - nearly 100% - of As(V). Changes in pH

had little effect on adsorption to halloysite indicating the potential formation of a

hydroxy-arsenate interlayer within the clay mineral structure. Chlorite also retained

arsenic well and was little influenced by pH. Its rich iron content may suggest arsenic

sorption onto iron oxide contained within its structure.

Manning and Goldberg (1997) conducted laboratory experiments on clay minerals to

determine:

a) the influence of pH and ionic strength (I) on As(III) adsorption;

b) differences between As(III) and As(V) adsorption;

c) stability of As(III) adsorbed onto clay mineral surfaces; and

d) to describe As(III) and As(V) adsorption envelopes using surface complexation

modelling.

They found that ionic strength had little effect on As(III) adsorption providing

macroscopic evidence for inner-sphere complexation. However, pH had a profound

influence on adsorption capability. All four materials (kaolinite, montmorillonite, illite

and synthetic amorphous Al(OH)3) exhibited low As(III) adsorption at low pH and

maximum As(III) adsorption between pH 7.5 and pH 9.5. This value is only slightly

above the first H3AsO30 pKa value of 9.2 and coincides with the point of zero charge

(PZC) for amorphous Al(OH)3. This suggests that Al-OH functional groups are highly

pH dependent and conducive to As(III) adsorption in preference to the SiO4 tetrahedra

predominant in the clay minerals utilised. Given this prediction, kaolinite, with its

dominant surface hydroxyl variable charge is expected to adsorb more As(III) than illite

which is controlled by permanent isomorphic substitution and minor surface edge ( Al-

OH and Al-OH2+) functional groups. The results were unexpected – illite adsorbed



more As(III) than kaolinite, and was postulated to be a result of the higher surface area

of illites, also noted by Lin and Puls (2000).

Desorption experiments (simulating mobilisation of arsenic) revealed significant

oxidation of As(III) to surface-bound As(V) had occurred. As(III) was stable in solution

at pH 4-9 but oxidation occurred at or above pH 9.2. Manning and Goldberg (1997)

suspected that the presence of MnO2 in the clays may have contributed to oxidation,

which in turn promotes more strongly bound arsenic in solid phase.

As(V) desorption was greatest at high pH, but not all arsenic was extractable. For

example, at pH 7 montmorillonite, kaolinite and illite desorbed As(V) at 80%, 66% and

33% respectively. Above pH 8.5 montmorillonite and kaolinite desorption was almost

100%, whereas illite retained more than 50% of its sorbed arsenic. The outcome of their

experiments suggested both pH and mineralogy were critical factors in controlling

arsenic adsorption to clay mineral surfaces.

Influence of Organic Matter

The presence of organic matter, both in solid and aqueous forms, can influence arsenic

sorption onto mineral surfaces. Organic matter often exhibits high surface areas and

great affinities for surface sorption. Humic substances act as good accumulators of

metal-humate complexes. Mukhopadhyay and Sanyal (2004) devised stability constants

for arsenic-humic/fulvic complexes that were considered stable. This was also

investigated by Saada et al. (2003) who found that at pH 7 humic substances adsorbed

onto kaolinite had a significant impact on arsenic adsorption. Redman et al. (2002)

found that the formation of aqueous complexes of arsenic and natural organic matter

dramatically reduced the amount of arsenic sorption onto hematite and displaced

already sorbed arsenic (III and V). Thus, the decomposition of organic matter can

release arsenic into solution (Belzile, 1988).

Additionally, the decomposition of other minerals by indirect organic matter influence

can affect arsenic mobility. The reductive dissolution of iron oxides catalysed by

organic matter oxidation is a prime example. Reduction of Fe(III) oxides may be

minimal if organic carbon contents are low, therefore producing an environment



favouring crystallisation (Willett and Walker, 1982). This is because organic matter acts

as a reductant, converting Fe(III) to Fe(II) and therefore inhibiting oxide formation, in

addition to inhibiting crystallisation (Schwertmann et al., 1968). If it is not present then

oxide formation and crystallisation will occur more easily. Willett and Walker (1982)

observed this process in the Shoalhaven River floodplain where cores in close proximity

to each other exhibited different iron oxide contents and degrees of crystallinity

correlating with organic matter presence. Crystallisation of iron oxides can enhance

arsenic release due to a reduction in surface area available for arsenic sorption. Thus the

presence of organic matter can directly and indirectly influence arsenic sorption

processes.

Carbonates

Arsenic incorporation within carbonate mineral structures has received little research

attention despite the ubiquitous nature of carbonate materials in the environment.

Arsenic partitioning on calcite surfaces is applicable to this investigation due to the

presence of calcite (shells) within the aquifer matrix. Cheng et al. (1999) provide one of

the few macroscopic investigations of arsenic association with calcite surfaces. Their X-

ray standing wave study concluded that the arsenite ion occupies the carbonate lattice

site of the calcite surface (Figure 3.6).

Actual bonding mechanisms remain unclear, however the authors suggest As(III) is

incorporated within the calcite lattice either at or near the surface (as opposed to ion

exchange reactions) potentially by a dissolution-reprecipitation mechanism.

Incorporation via an outer-sphere complex could suggest arsenic adsorption in alkaline

environments given that the PZC of calcite ranges from 7-10.8 (Romero et al., 2004).

Goldberg and Glaubig (1988) observed a maximum surface adsorption of As on calcite

at around pH 10. Subsequently, Romero et al. (2004) conducted batch experiments on

carbonate-rich aquifer material from Zimapan, Mexico, and found arsenic adsorption

was greatest (97%) at pH 10.5. These results indicate that arsenic adsorption onto

calcite is likely to exert a major control on dissolved arsenic concentrations in the

groundwater. Goethite was also found to influence arsenic sorption in this study,

however, calcite sorption was found to dominate by using a combination of

experimental (determination of PZC; batch experiments; pH dependent sorption) and



geochemical modeling (MINTEQA2) techniques. The high affinity of arsenic for the

carbonate-rich aquifer material was suggested to be utilised as a remediation option by

allowing deeper contaminated groundwaters to flow through shallow carbonate-rich

material in order to decrease dissolved arsenic concentrations.


Figure 3.6 As(III) substitution for the carbonate ion on the calcite surface as suggested by Cheng et al. (1999).

Sulfides other than Iron

Arsenic association with iron sulfides has been touched briefly upon in the discussion

on arsenian pyrite. Other sulfides, however, possess the potential for anion retardation

via surface complexation mechanisms. Bostick et al., (2003) studied arsenite adsorption

on the more stable sulfide ores found in anoxic environments; galena (PbS) and

sphalerite (ZnS). They found a distinct difference between arsenic adsorption on

sulfides when compared to arsenic adsorption processes postulated for iron oxide

surfaces. The main notable difference was an increase in As(III) adsorption with

increasing pH on PbS and ZnS. This is in direct contrast to arsenic-oxide adsorption

which is greatest at circumneutral pH and lowest at high pH. These results suggest

arsenic adsorption on sulfide minerals does not occur via ligand-exchange mechanisms.

Bostick et al. (2003) also ruled out isomorphic susbstitution with Pb or S. Instead they

proposed the formation of an inner-sphere multinuclear arsenic thiosulfide complex on

the surface of PbS and ZnS.



3.4.3.2.2 SURFACE PRECIPITATES

Surface precipitates were briefly mentioned in the discussion on arsenic association

with pyrite. As surface coverage of adsorbed anions increases, three dimensional

surface precipitates can form (Sparks, 2003). The formation of metal hydroxide surface

precipitates may be an important mechanism for retaining metals in the solid phase,

thereby limiting transport and bioavailability. Surface precipitation may take place at

high As(III) activity (Bostick et al., 2003). In sulfidic, reducing environments

precipitation of arsenic sulfides such as orpiment (As2S3) have been proposed for

mechanisms of arsenic retention (Sadiq, 1990). Bostick et al. (2003) conducted

laboratory experiments and concluded it was unlikely for arsenic sulfide precipitation to

occur on galena and sphalerite but did emphasize the need for additional research

involving such reactions.

3.4.3.2.3 COMPETITIVE ANION EXCHANGE

The mobility, bioavailability and toxicity of arsenic may be influenced by the

competitive interaction of other anions. Various minerals are affected differently and the

concentrations and speciation of competitive anions can also influence potential for

arsenic sorption. Common competing anions include SO42-, PO4

3-, CO32- and Cl-.

Perhaps the most commonly studied competitive anion with arsenic is PO43-, probably

due to its application via fertilizer use increasing its potential to be present in

groundwater systems. Competitive ligand exchange generally occurs between H2PO4-

and HPO42- and arsenate ions (Manning and Goldberg, 1997).

Aside from fertilizer use, Belzile (1988) suggested anion exchange processes were

responsible for occurrence of dissolved arsenic in porewaters of offshore sediments.

Phosphate was suggested to replace arsenic in pyrite (equation 3.6) maintaining

dissolved arsenic concentrations while depleting dissolved phosphate in the porewaters:

PO43- + As-FeS2 As5+ + PO4-FeS2 (3.6)



Goh and Lim (2004) studied anion competition with arsenic sorption on a bulk tropical

soil sample as opposed to singular or synthetic minerals. They found that SO42- rarely

affected As(V) adsorption capacity but could compete with As(III) for sorption sites.

Phosphate had a more profound effect on sorption capacity for both As(III) and As(V),

which may be attributed to its higher negative charge. Goh and Lim (2004) suggested

two possible reactions for PO43- on surface sites; strong inner-sphere complexation

and/or accumulation or precipitation of PO43- providing electrostatic repulsion on the

mineral surface. Both would reduce arsenic sorption and increase its mobilisation into

groundwater.

Both As(III) and As(V) adsorption onto goethite is reduced in the presence of dissolved

organic matter (Grafe et al., 2001). Dissolved organic matter (DOM), such as humic or

fulvic acids, form stable complexes with mineral surfaces (Kaiser et al., 1997) thereby

blocking arsenic sorption sites on iron oxides, alumina, kaolinite or quartz (Grafe et al.,

2001). This has recently been supported by Bauer and Blodau (2005) who confirmed

competition between arsenic and organic anions for sorption sites on synthetic iron

oxides and natural soils and sediments. Therefore DOM may contribute to increasing

the mobility and potential bioavailability of arsenic via competitive anion exchange

mechanisms.

Appelo et al. (2002) examined ferrihydrite sorption capacity and found that carbonate

present in common soil and groundwater concentrations significantly replaced arsenic

on sorption sites. This has been further confirmed by Anawar et al. (2004) who

conducted batch experiments that showed bicarbonate solutions were effective in

releasing arsenic from iron and manganese oxyhydroxides. The formation of arseno-

carbonate complexes; As(CO3)2-, As(CO3)(OH)2- and AsCO3

+, which are stable in water,

were earlier predicted by Kim et al. (2000).

Manning and Goldberg (1996) found that molybdate (MoO42-) inhibited As(V)

adsorption onto oxide minerals only at pH levels below 4. Similarly, Ronngren et al.

(1994) suggested sulfide may inhibit As(III) adsorption onto mineral surfaces due to

competition for surface sites. In general agreement with the studies cited above, Goh



and Lim (2005) found the ability of anions to mobilise arsenic from a fine soil followed

the order PO43- >> CO3

2- > SO42- and Cl-.

3.5 MOBILISATION OF ARSENIC: THE GEOCHEMICAL TRIGGERS

The aqueous distribution of many trace elements is often dominated by solution pH;

redox potential; composition and interaction with the surrounding solution; and surface

reactions on solid particles (both colloids and in-situ matrices). This current section

explains geochemical conditions required for the mobilisation of arsenic. Colloidal

interactions are included here as many colloids pass through a 0.45 m filter (utilised

herein) and are thus included in the ‘dissolved’ analyses; their fractionation cannot be

determined otherwise with the use of the sampling methods employed herein (see

Appendix A for sampling methodology).

3.5.1 Changes to Aqueous pH As noted in the previous section, solution pH influences the adsorption of both As(III)

and As(V). At slightly acidic to neutral pH values commonly encountered in most

natural waters, arsenic is strongly adsorbed by oxide minerals. As the pH approaches 9

desorption of arsenic can occur as mineral surfaces become increasingly negatively

charged, repelling the As oxyanion. Smedley and Kinniburgh (2002) attribute

desorption at high pH as being one of the most likely mechanisms for high arsenic

groundwaters under oxidising conditions. Many authors (see Robertson, 1994) note a

positive correlation between dissolved arsenic concentrations and pH in predominantly

oxidising conditions. The oxyanions of Se, Cr, Mo and V generally correlate positively

with pH and As under these conditions, indicating similar pH-influenced mechanisms

governing their mobility.

Changes in aqueous pH may also influence mineral solubility. Based on the literature

reviewed for carbonate minerals, uncertainty remains over the association of arsenic

with minerals such as calcite, siderite and dolomite; but the inclusion of arsenic in the

calcite structure discussed previously indicates possible release of arsenic upon calcite

dissolution (equation 3.7).



2Ca(As,CO3) + 4H+ 2Ca2+ + As +2CO2 + 2H2O (3.7)

3.5.2 Shifts in Aqueous Redox Potential Following general redox trends in natural systems, several stages of the redox cycle can

influence arsenic mobilisation processes. These are described in (Table 3.10) with

specific examples of arsenic release or retardation processes provided in response to

changes in aqueous Eh.

Dissolved O2 is consumed first while dissolved CO2 increases via the decomposition of

organic matter. Arsenic bound to organic matter may be released upon its degradation,

such as arsenic associated with coffee rock in shallow groundwaters where oxygen is

present. The oxidation of arsenic-bearing minerals by oxygen can also release arsenic

into dissolved phase, particularly arsenic associated with sulfide minerals like arsenian

pyrite. The oxidation of sulfides, however, is not an efficient mechanism for arsenic

release into groundwater due to its rapid adsorption onto iron oxide by-products

commonly formed upon sulfide oxidation.

When oxygen is consumed nitrate becomes the dominant electron donor and can be

reduced by oxidising arsenian pyrite. Manganese reduction generally follows,

increasing dissolved Mn2+ concentrations. Arsenic sorbed onto Mn oxides may be

released as a consequence. Iron oxides are similarly reduced increasing dissolved Fe2+

in solution. In strongly reducing conditions, Fe(III) on the surface of iron oxides reduces

to Fe(II), which lowers the overall net positive charge of the oxide producing weaker

electrostatic bonds between sorbed anions and the oxide surface, leading to desorption

of the anion. Iron oxide dissolution occurs under strongly acid and strongly reducing

conditions. The actual dissolution of iron oxides will release both labile and non-labile

arsenic (and any other bound trace elements) into solution as the mineral structure

completely breaks down.

Tab

le 3

.10

Cha

nges

to r

edox

equ

ilibr

ium

in n

atur

al w

ater

s and

how

this

may

aff

ect a

rsen

ic m

obili

satio

n.

Indi

cativ

eEh

Elec

tron

Acc

epto

rEx

ampl

e Pr

oces

s R

elea

sing

A

rsen

icEq

uatio

nEx

ampl

eR

efer

ence

Slig

htly

oxid

isin

gC

onsu

mpt

ion

of o

xyge

n

Dec

ompo

sitio

n of

org

anic

mat

ter

and

rele

ase

of b

ound

ars

enic

Oxi

datio

n of

ars

enia

n py

rite,

w

ith a

dsor

ptio

n on

to ir

on o

xide

by-

prod

ucts

Hum

ic a

cid

As

+ 2O

2 + H

+

2CO

2 + 2

H2O

+ A

s (aq

)

Fe(A

s,S

) 2 +

15/4

O2 +

7/2

H2O

Fe(O

H) 3

As

+SO

42-+

4H+

Muk

hopa

dhya

yan

d S

anya

l (20

04)

Kim

et a

l. (2

002)

Slig

htly

redu

cing

Nitr

ate

redu

ctio

nN

itrat

e re

duct

ion

by p

yrite

oxi

datio

n8F

e(A

s,S

) 2 +

13N

O3- +

25H

2O +

10H

+

8Fe2+

+ 8

HA

sO42-

+ 8

SO

42- +

13N

H4+

App

elo

and

Pos

tma

(199

9)

Slig

htly

redu

cing

Man

gane

sere

duct

ion

Man

gane

se o

xide

redu

ctio

n an

d re

leas

e of

sor

bed

arse

nic

MnO

2H

AsO

2 + 2

H+

H3A

sO4(

aq) +

Mn2+

Cha

illou

et a

l. (2

003)

Red

ucin

gIro

nre

duct

ion

Iron

redu

ctio

n on

oxi

de s

urfa

ces

chan

ges

net c

harg

e le

adin

g to

ar

seni

c de

sorp

tion

Iron

oxid

e re

duct

ive

diss

olut

ion

and

rele

ase

of s

orbe

d ar

seni

c ca

taly

sed

by o

rgan

ic m

atte

r deg

rada

tion

FeA

sO4- +

2e- +

5H

+

Fe2+

+ H

3AsO

3 + H

2O

8FeO

OH

As

+ C

H3C

OO

H +

14H

2CO

38F

e2+ +

16H

CO

3- +12

H2O

+ A

s (aq

)

Nic

kson

et a

l. (2

000)

Nic

kson

et a

l. (2

000)

Red

ucin

g A

rsen

ate

redu

ctio

nA

s(V

) red

uctio

n to

As(

III) a

nd

chan

ge in

sor

ptio

n pr

oper

ties

AsO

43- +

2H

+ + 2

e- A

sO33-

+ H

2OB

ose

and

Sha

rma

(200

2)S

trong

lyre

duci

ngS

ulfa

tere

duct

ion

Pre

cipi

tatio

n of

inso

lubl

e su

lfide

s an

d im

mob

ilisa

tion

of a

rsen

ic

Fe2O

3 + 4

SO

42- +

8C

H2O

+ 1

/2O

2 + A

s (aq

)2F

e(A

s,S

) 2 +

8H

CO

3- + 4

H2O

Zhen

g et

al.

(200

4)



The full reaction process of reductive dissolution of iron oxides remains unclear,

however, it is generally responsible for the high amounts of Fe(II) in reducing waters

(0.1 – 30 mg L-1). Smedley and Kinniburgh (2002) state that it does not appear to be

responsible for all dissolved arsenic in arsenic-rich reducing groundwaters. These

strongly reducing conditions are often brought about by rapid sediment burial and can

be dependent on the organic content of the sediments, which drives the redox reaction.

Arsenic release by reduction of iron oxides due to the decomposition of organic matter

has been linked to the release of arsenic in some groundwater environments (McArthur

et al., 2004).

Reduction of As(V) to its less strongly sorbed counterpart As(III) may increase the

mobility of arsenic in the environment via decreased affinity for surface adsorption.

Redman et al. (2002) and more recently Bauer and Blodau (2005), found that DOM is

capable of reducing arsenate and oxidising arsenite; both redox changes having an effect

on arsenic mobility. Arsenite oxidation can also be catalysed by dissolved manganese.

Next sulfate is converted to S2- which subsequently reacts with any dissolved Fe2+ to

form insoluble precipitates which may sequester arsenic. Sulfide is normally the

limiting ion for iron sulfide mineral formation, which leads to an excess of dissolved

Fe2+ over dissolved S2-. This is particularly important in low sulfate environments.

Changes in Eh equilibrium described in Table 3.10 can thus lead to the dominance of

one mechanism over another, or conditions can be reversed and redox cycling may be

observed whereby arsenic is repeatedly mobilised and immobilised as the Eh conditions

change. An indication of a change in redox potential occurring in sediments can be

demonstrated by a colour change from red/brown/orange (oxidising conditions) to

grey/green/blue (reducing conditions). Additionally, many of these processes are

mediated by biological reactions rather than simple chemical reactions alone.

3.5.3 Influence of, and Interaction with, the Surrounding Solution The immobilisation of As via competitive anion exchange (Section 3.4.3.2.3) can

generally be reversed to mobilise As into solution once again. This has previously been



explained, therefore discussion moves onto aqueous complexation mechanisms.

Complex formation is defined loosely by Morel and Hering (1993) as,

“the reversible reaction of two dissolved species to form a third one” (pg 319)

and generally takes place between metals of one charge and ions of the opposite charge

in aqueous solution. The ion bonding to the metal is termed a ligand. Common ligands

include Cl-, Fl-, Br-, NO3-, CO3

2-, SO42-, NH3, S2-, PO4

3- and a variety of organic

molecules with suitable surface charges. Complexation can inhibit arsenic adsorption

thereby enhancing its mobility in aqueous environments.

Cullen and Reimer (1989) reviewed detailed calculations of the equilibrium speciation

of numerous trace elements and found that, unlike other elements, few aqueous arsenic

complexes form with typical water constituents. Arsenic-fluoride complexes were

possibly suggested at high fluoride concentrations, and in seawater arsenate may pair

with magnesium and calcium. Apart from these they concluded that the range of water-

soluble arsenic complexes was limited. More recent research has revealed the formation

of stable As(III) sulfide solution complexes which have been postulated to enhance the

mobility of As(III) by maintaining it in a soluble form (Helz et al., 1995; Rochette et al.,

2000). Such complexes include AsO(SH)2- in slightly sulfidic solutions. Bostick et al.

(2003) considered that the thioarsenite trimer As3S3(SH)3 in higher sulfide activities

may be a plausible complex present in their experimental results.

The formation of aqueous complexes of arsenic with DOM has recently been examined

by Saada et al. (2003) and found to increase arsenic mobility in waters containing

DOM. Wetland soils, which often contain DOM at slightly higher than normal

concentrations, have shown an increase in arsenic mobility in the presence of high DOC

concentrations (Kalbitz and Wennrich, 1998).

The formation of aqueous complexes is also important when examining the

bioavailability of metals. Some metals are much less toxic when strongly complexed to

organic ligands compared to when present as free ions or hydroxylated (Drever, 1997).

Others are more conducive to plant uptake when complexed, such as cadmium

complexed with chloride.



3.5.4 Colloidal Transport Groundwater samples are often filtered through a 0.45 m filter to remove suspended

sediments and algae before chemical analysis. The size range of many common colloids

means that some will pass through this filter size therefore giving the appearance of

being contained within the dissolved fraction of a water sample. Their small size means

they exhibit large surface areas prone to sorbing ions from solution. These ions are thus

reported as being present in the dissolved fraction rather than specifically adsorbed to a

solid colloidal phase. This can have implications for predictions on the bioavailability of

trace elements and lead to the increased transport of adsorbed elements through an

aquifer. Sorption reactions pertaining to solid phase aquifer matrices are also applicable

to colloids. Thus, when the pH of the ambient solution is below the PZC (Table 3.11) of

the colloid the surface becomes protonated resulting in a net positive charge.

Table 3.11 Common colloids and their PZC (vanLoon and Duffy, 2000). Colloid PZC

SiO2 2.0 MnO2 2.0 – 4.5

Fe2O3 hydrated 6.5 – 9.0 Goethite 7.5

Haematite 8.5 Al2O3 hydrated 5.0 – 9.0 Humic material 4.0 – 5.0

Bacteria 2.0 – 3.0

Anions will sorb to the colloid surface via electrostatic interaction. Many of the colloids

listed below have already been shown to sequester arsenic by surface complexation

reactions. Colloidal iron oxides are most likely to sorb As under natural pH conditions

and are common forms of As transported in river waters derived from hinterland iron

oxide and arsenic sources.

3.6 NATURALLY OCCURRING GROUNDWATER ARSENIC

With the increased testing of trace elements in groundwater around the globe, high

arsenic occurrences have been detected in many aquifers used for drinking water

purposes. Smedley and Kinniburgh (2002) provided a review of known arsenic-rich



aquifers and their status, current to several years ago, producing the global distribution

diagram shown in Figure 3.7. Conditions producing these elevated arsenic

concentrations are further reviewed and built on with current literature below. In

particular, the status of arsenic in the Australian environment is discussed in light of its

previous omission in worldwide occurrences.


Figure 3.7 Natural arsenic distribution in aquifers globally (Smedley and Kinniburgh, 2002). Some mining and geothermal sources are also noted.

Naturally occurring arsenic-enriched aquifers are common throughout many countries

but they are not ubiquitous. Elevated arsenic in groundwater is related to specific

geochemical conditions and the past and present hydrogeology of the aquifer. A variety

of case studies from different countries, geochemical conditions, and past depositional

environments, are presented to indicate the heterogeneity and widespread global

occurrence of elevated arsenic in groundwater.



3.6.1 Geochemical Controls

3.6.1.1 Oxidising Environments Oxidising conditions are generally conducive to As(V) domination in aqueous

environments. The common association of arsenic with many sulfide minerals means

that oxidising conditions in groundwater elevated in arsenic is often derived from the

oxidation of such minerals. The sporadic occurrence of arsenic in the oxidising

conditions of the Goose River basin in Maine has been investigated by Sidle et al.

(2001); Sidle (2002, 2003) and Sidle and Fischer (2003). Arsenic is predominantly

present as As(V) in both H2AsO4- and HAsO4

2- states. The slightly oxidising conditions

present in the aquifer (median dissolved oxygen [DO] 8.1 mg L-1) promotes oxidation of

arsenian pyrite along fractured zones in the igneous and metamorphic bedrock. Sidle et

al. (2001) suggest these arsenian pyrites are the probable source of arsenic

concentrations in groundwater. This suggestion is supported by analysis of 34S and 18OSO4

/18OH2O (Sidle, 2002) isotopes which indicate SO42- presence with elevated As in

solution may be characteristic of arsenian pyrite oxidation in the fractured flow systems.

In the confined part of the central Oklahoma aquifer dissolved arsenic concentrations

reach 110 g L-1 under oxidising conditions (Schlottmann and Breit, 1992). As(V) was

determined to be present and strongly dependent on pH. The authors found that the

highest reported arsenic concentrations were only observed at pH values above 8.5 and

could be attributed to desorption from abundant iron oxides present in the aquifer.

An unusual occurrence of arsenic release in an oxidising environment occurs in a

confined aquifer in eastern Wisconsin. The fully saturated nature of deeper

groundwaters or confined aquifers often leads to reducing conditions where sulfide

mineral oxidation does not readily occur. The Fox River Valley, however, reports

arsenic up to 12,000 g L-1 where DO is typically less than 0.5 mg L-1 (Schreiber et al.,

2000). Sulfur isotopic signatures suggest oxidation of arsenian sulfide as the source of

arsenic to groundwater. Four potential oxidant sources were suggested (regional

recharge, vertical leakage, borehole and dewatering), however, interaction between

oxygen in the borehole and the arsenian pyrite horizon was implied as the probable

oxidant source.



3.6.1.2 Arid Oxidising Environments The adsorption characteristics of As(V) are such that it becomes less strongly bound at

increasing pH levels, notably pH 8.5 – 9.5. Arid conditions often promote pH values in

this range thus leading to the mobilisation of arsenic in groundwater. Desorption of

arsenic from metal oxides is considered an important mobilisation process under these

geochemical conditions. If these conditions are widespread then arsenic contamination

may also be widespread. Such conditions are prevalent in Chaco-Pampean aquifers of

Argentina. Arsenic-related health problems have been recorded in this region since 1917

(Smedley et al., 2002).

Robertson (1994) found arsenic present entirely as As(V) in the HAsO42- state in closed

alluvial basins of Arizona. The dominant control mechanism for arsenic in the

groundwater system was assumed to be the sorption or desorption of HAsO42- on iron

oxide surfaces. This process is promoted by the increase in pH along flowpath as

silicate weathering releases H+ ions into solution. Large concentrations of arsenic were

often reported in the central portions of the basin where conditions became suitable for

arsenic desorption under increased pH conditions. The observed correlation between

arsenic and other oxyanions in solution (Mo, Se, V) suggests competition for adsorption

sites is also occurring in these alluvial basins.

Where evaporation rates exceed precipitation arsenic may become concentrated in many

hydrologically closed basins. Areas affected by evaporative concentration of arsenic

include many parts of western USA (Welch et al., 2000), Argentina, Mexico and Chile

(Caceres et al., 1992). High pH values of saline waters promotes desorption of arsenic

and it does not appear to easily partition into evaporative minerals, leading to high

concentrations in dissolved phase (Welch et al., 2000). The Owens Lake in California

provides an excellent example of evaporative concentration of arsenic. Dissolved As

concentrations can be as high as 96 mg L-1 (Ryu et al., 2002). Levy et al. (1999) found

As and F behaved conservatively during the onset of salinity; remaining in solution

during the precipitation of carbonates, halite and thenardite. This causes elevated

concentrations of both elements in shallow groundwaters of the dry lake-bed with each

exhibiting an extremely strong correlation (r2 > 0.94) with the evapoconcentration



index, lithium. Arsenic and F are not incorporated into salts until very high salinities are

reached (> 9 molar) but can be easily released during rainfall events (Levy et al., 1999).

In the semi-arid Muzaffargarh District of Pakistan, Nickson et al. (2005) predict

dissolved As derived from evaporative concentration by canal irrigation could

potentially reach mg L-1 levels but in fact remain less than 25 g L-1. They attribute

these low concentrations to sorption of As from the oxic groundwater to aquifer

sediments.

3.6.1.3 Reducing Conditions The occurrence of high As groundwaters under reducing conditions has been

documented in several parts of the globe. The most notable (in terms of people at risk)

occurrence of arsenic under these conditions occurs in the alluvial aquifers of

Bangladesh. There are several dominant arsenic mobilisation theories that have been

proposed for Bangladesh.

The oxidation theory involves arsenic release during pyrite oxidation induced by

lowering of the water table during groundwater extraction (Mallick and Rajgopal,

1996). This theory has largely been discounted as a dominant mobilisation process

given the highly reducing conditions observed in the aquifer.

The impact of fertilizer-derived phosphate addition to the aquifer was predicted to

release arsenic sorbed to minerals in the aquifer matrix due to competitive anion

exchange (Acharyya et al., 1999). However, phosphate transport to depths in the aquifer

is unlikely, which does not account for the high concentrations of arsenic observed in

groundwater at depth.

As such, the most commonly accepted theory for arsenic release in Bangladesh

groundwaters is attributed to the reduction of iron oxides and release of sorbed arsenic

(Nickson et al., 2000; Anawar et al., 2003; Tareq et al., 2003). Various trace elements,

particularly arsenic, are known to be scavenged by oxides of Fe and Mn. When these

oxides are subjected to a reducing environment they can be dissolved along with any

adsorbed trace elements. The actual mechanism of arsenic release in Bangladesh is not



clear, however Smedley and Kinniburgh (2002) suggest that a combination of reductive

dissolution of iron oxides and desorption from various oxide and clay mineral surfaces

could be responsible for the elevated concentrations in groundwater.

More recent investigations in Bangladesh have focussed on the association of organic

matter with arsenic release in anoxic groundwater (McArthur et al., 2004; Nickson et

al., 2000; Harvey et al., 2002). These investigations attempt to define the process

driving the reduction of iron oxides and subsequent release of arsenic. The source of

organic matter in the aquifers is under debate. Harvey et al. (2002) suggest organic

matter from surface water is drawn into the aquifer during irrigation pumping which

promotes reduction and arsenic release. This has been most recently disputed by

McArthur et al. (2004) who found natural organic matter is present in peaty strata

throughout the aquifer and that this is the probable source driving reduction (also

proposed by McArthur et al., 2001; Ravenscroft et al., 2001). An additional finding of

the 2004 study by McArthur et al. is that faecally-derived organic matter probably is

unlikely to cause the arsenic release.

3.6.1.4 Combined Oxidising and Reducing Conditions Variation in redox potential can be observed within an aquifer. The most obvious

variation is a decrease in oxygen with depth in an unconfined aquifer resulting in a

change from oxidising to reducing conditions. This redox variation has been observed in

aquifers in Ghana (Smedley, 1996). Shallow waters in the heavily leached regolith

overlying bedrock are oxidising, promoting oxidation of arsenic sulfide minerals and

sorption onto iron oxides in the acidic-generated environment. Slightly deeper waters

become more reducing due to consumption of DO during oxidation in combination with

increased pH from water-rock interaction, which decreases arsenic sorption onto iron

oxides thereby increasing its concentration in dissolved phase. The deepest

groundwaters are highly reducing with few electron acceptors to drive sulfide mineral

oxidation keeping dissolved arsenic concentrations relatively low.

Arsenic cycling was observed by Belzile (1988) in sediments of the Laurentian Trough.

Pyrite oxidation contributed dissolved arsenic to the system which was then scavenged

by iron oxides. The onset of reducing conditions allowed release of arsenic that was



reprecipitated into pyrite once more. This cycle occurred through the sediment-water

interface to reducing depths several metres below.

More recently, Zheng et al. (2004) suggested that re-oxidation of reducing As and Fe

enriched water in Bangladeshi aquifers was plausible. Sulfur isotopic ratios showed that

conditions were once reducing enough to mobilise arsenic, but the concurrent presence

of O2 and tritium suggested recent exposure to the atmosphere, producing alternative As

mobilisation processes in a redox-controlled cyclical manner.

3.6.2 Geological and Depositional Influences

3.6.2.1 Geothermal Systems Arsenic-rich geothermal waters have been reported in the USA (Welch et al., 1988,

2000), New Zealand (Robinson et al., 1995), Japan (Yokoyama et al., 1993), Iceland

and Kamchatka (White et al., 1963). The high temperatures associated with geothermal

waters can be conducive to mineral dissolution releasing trace elements into dissolved

phase. Several authors (Welch et al., 1988; Wilkie and Hering, 1998) report a

correlation between geothermal arsenic and saline waters. Arsenic may also be

correlated with temperature as reported by Gemici and Tarcan (2004) in geothermal

waters used for swimming and spa facilities in Turkey. Concentrations of dissolved

arsenic reach 1,417 g L-1 in these thermal waters, however, Ellis and Mahon (1977)

report concentrations in the high range of 45,000 – 50,000 g L-1 for geothermal waters

from the Antofagasta region of Chile.

3.6.2.2 Volcanic Sediments The metavolcanic deposits forming Bowen Island in British Columbia contain arsenic-

bearing sulfide mineralisation which have been oxidised, allowing dissolved arsenic to

enter the groundwater system (Boyle et al., 1998). The Ischia Island in Southern Italy is

formed from Quaternary volcanic rocks and exhibits intense hydrothermal activity;

Daniele (2004) found groundwater arsenic to be volcanic in origin but intensified by

geothermal activity. A volcanic-sedimentary sequence of rocks in the Zimapan Valley,

Mexico, provide the host sediment for basaltic dykes. These dykes caused



metamorphism during emplacement, forming arsenic sulfides (Rodriguez et al., 2004)

which have subsequently caused elevated arsenic in the groundwater.

3.6.2.3 Alluvial and Deltaic Sediments Unconsolidated aquifers can cover large areas, often in heavily populated coastal zones,

and are capable of storing good quality groundwater resources. It is unfortunate then

that many of these productive aquifers are prone to natural arsenic contamination. The

most severe case, in terms of population exposed to arsenic rich groundwater, has

previously been discussed (Bangladesh). The Bengal Delta Plain consists of Holocene

alluvium derived from the Himalayan and Indo-Burman range and exhibiting dissolved

arsenic concentrations up to 2,500 g L-1 in the aquifer (BGS and DPHE, 2001). The

Red River Delta in northern Vietnam has only recently (1998) discovered arsenic up to

3,050 g L-1 in its groundwater (Berg et al., 2003). The similarity of geology and

groundwater composition between the Bengal Delta and the Red River Basin were

predicted, and confirmed, to exhibit elevated arsenic concentrations in the alluvial

aquifers surrounding Hanoi (Berg et al., 2001). Additionally in Nepal, alluvial aquifers

contain up to 409 g L-1 of dissolved arsenic (Bhattacharya et al., 2003).

3.6.2.4 Consolidated Sediments The incidence of arsenic contamination in unconsolidated sediments is perhaps more

pronounced given their capacity to provide large quantities of water to many people. In

spite of their less frequent use for domestic water supplies globally, consolidated

aquifers are important resources and can be heavily exploited for their groundwater.

Like the unconsolidated aquifers, these water resources have not been left untouched by

natural arsenic contamination. Examples are provided below for sandstone, limestone

and shale consolidated aquifers.

In an area where no obvious source of arsenic occurs, the Triassic Keuper Sandstone

aquifer in northern Bavaria exhibits arsenic concentrations up to 150 g L-1. Heinrichs

and Udluft (1999) found the arsenic was related to the geology and lithofacies of the

aquifer. Terrestrially derived sediments of fluvial origin were responsible for arsenic

concentrations in comparison to adjacent marine beds that contained no arsenic. In a



separate study on arsenic occurrence in a sandstone aquifer of the Fox River Valley,

Thornburg and Sahai (2004) found that where discrete zones of arsenic rich sulfides

were lacking within the aquifer matrix, arsenic was present as disseminated veins,

grains and nodules. These discrete sulfides were found to influence groundwater arsenic

concentrations sporadically throughout the sandstone and dolomite aquifer.

The mining town of Zimapan, Mexico, is well known for its base-metal sulfide

mineralisation which intrudes domestic water supply limestone aquifers. Arsenic

concentrations reach more than 1 mg L-1 in the most polluted drinking water wells

installed in carbonate-rich aquifer material (Romero et al., 2004). The arsenic present in

this limestone aquifer is directly related to a nearby mineralised source rather than

dominated by aquifer lithology. Limestone aquifers generally do not contain naturally

elevated arsenic (see Al-Awadi et al., 2003, for an example of a limestone aquifer

without elevated arsenic) unless some association with a proximal arsenic source is

recognised.

Examples pertaining to shale aquifers are rare, potentially due to their low water holding

capacity and subsequent rejection for domestic yield requirements. Chen et al. (1994)

linked the occurrence of blackfoot disease with arsenic concentrations in groundwater

extracted from sands, muds and shale in Taiwan. However, the characterisation of

arsenic distribution in this aquifer remains poorly understood. Average arsenic

concentrations in shale rock assays are generally higher than other rock types indicating

groundwater in shale aquifers may be influenced by elevated As. More research is

needed to fully characterise the association between shale aquifers and arsenic

occurrence.

3.6.2.5 Lacustrine Environments Shallow groundwater in the Carson Desert, Nevada consists of elevated arsenic and

uranium concentrations controlled by evaporation, adsorption and redox processes

(Welch and Lico, 1998). The aquifer formation consists of deep-lake sediments, lake-

bed deposits and alluvium but is largely controlled chemically by the above processes.

The authors postulated it may be a modern analog of non-marine closed basins

producing U-ore deposits in New Mexico.



The Huhhot Basin, Inner Mongolia, while also of (fluvio)lacustrine origin is dominated

by anaerobic conditions with elevated organic acids in addition to elevated arsenic

(Smedley et al., 2001). The reducing conditions were considered similar to those

observed in Bangladesh groundwaters. Association between arsenic and organic acid

was suggested to be a product of the highly reducing conditions or potential competition

between sorbing anions. Once again, geochemical conditions rather than

sedimentological controls influence arsenic occurrence in this fluviolacustrine

environment.

3.6.2.6 Glacial Drift The erosion of arsenic-bearing rocks by glacial processes can result in complex arsenic

distributions in groundwater related to the heterogenous nature of glacial aquifer

deposits. A deep glacial drift aquifer consisting of sand and gravel is the primary

groundwater drinking source in the lower Illinois River Basin (Warner, 2001). As part

of the USGS National Water Quality Assessment Program, arsenic concentrations were

investigated and found to be positively correlated with barium and chloride. A

subsequent increase in concentrations of these analytes along the flowpath indicates

arsenic may be derived from the underlying pyrite-containing bedrock, which

discharges into the glacial drift aquifer through localised structural features (Warner,

2001). Given the common occurrence of glacial drift aquifers overlying shale and coal

bedrock in the Midwest, in combination with the complex geochemistry of arsenic

resulting in variable distribution throughout the aquifer, arsenic is considered a health

concern for the entire Midwest region of the United States.

Szramek et al. (2004) also note the complexity associated with arsenic geochemistry in

glacial drift aquifers as a product of past glacial history, erosion, deposition and aquifer

recharge rate fluctuations. Different watersheds exhibited variation in arsenic

concentrations attributable to aquifer heterogeneity and consistent with groundwater

residence times. That is, less permeable aquifers reported higher arsenic concentrations.

The original source of arsenic is derived from the erosion and oxidation of arsenopyrite

during glacial processes.



In addition to providing valuable groundwater resources, glacial processes can deposit

low permeability till forming confining units/aquitards. These tills are capable of

retarding dissolved solutes like arsenic, thereby removing it from solution (Yan et al.,

2000). Overall, glacial drift aquifers can provide complex arsenic occurrences related to

aquifer heterogeneity, which may include geochemical processes conducive to the

addition and removal of arsenic to/from the groundwater resource.

3.6.2.7 Aeolian Loess Deposits In contrast to glacial aquifer deposits, aquifers composed of loess are lithologically

homogenous. Therefore, variation in groundwater chemistry can be attributable to

hydrochemical differences rather than lithological controls. The La Pampa Province of

central Argentina has significant groundwater problems due to high concentrations of

various oxyanions, including arsenic (Bundschuh et al., 2004; Farias et al., 2003;

Smedley et al., 1998, 2002). The source of arsenic is thought to be derived from the

weathering of primary silicate minerals contained within fine-grained loess and ash

deposits forming the main exploited aquifers of the region (Smedley et al., 1998).

Geochemical processes are the controlling factors in arsenic mobilisation and are

characterised by arid oxidising conditions. In addition, slow groundwater flow rates

through the fine-grained material have enabled the accumulation of arsenic within the

aquifer and decreased its opportunity for removal via aquifer flushing (Smedley et al.,

2002).

3.6.2.8 Zone of Water Table Fluctuation The vertical movement of the water table can lead to chemical changes in the

unsaturated and shallow saturated zone. For example, eustatic changes in sea level can

enable flocculation and precipitation of humus leaching through aquifer materials

leading to impregnation of layers often referred to as coffee rock, waterloo rock, coastal

sandrock or indurated sand. The presence of these organic layers can influence arsenic

mobilisation via competitive anion exchange with dissolved organic matter or reduction

of oxides in the presence of organic matter subsequently releasing arsenic into solution

(Nickson et al., 2000).



The oxidation of pyrite is a common occurrence leading to the development of ASS in

low-lying coastal areas. Oxidation often occurs as a product of vertical movement of the

water table. Arsenian pyrite oxidation can contribute dissolved arsenic to the

groundwater system if it is not sequestered by other processes. Iron oxide phases

commonly form as by products of pyrite oxidation and are excellent scavengers of

arsenic. Dudas (1984) found that iron oxides were favoured as arsenic scavengers over

the inclusion of arsenate for sulfate in the jarosite (FeSO4) structure; another commonly

formed pyrite oxidation by-product. Arsenic correlation with ASS may be a problem for

many coastal areas (O’Shea and Jankowski, 2001). Appleyard et al. (2004) found that

water table decline in a coastal suburb of Perth, Australia, led to the oxidation of

sulfidic peat soils resulting in widespread acidification and contamination of

groundwater by arsenic.

Some geochemical processes may be enhanced at the zone of water table fluctuation.

Smedley et al. (2002) noted more arsenic release at the water table compared to greater

depths in the aquifer studied. No correlations existed to indicate evaporation, and as

such, desorption from metal oxides and silicate reactions were suggested as the

controlling arsenic mobilisation factors. Water table fluctuations under irrigated

agricultural land can alter natural hydrodynamics and chemical processes. Northey et al.

(in press) observed this phenomenon in relation to major ion chemistry, however trace

elements may also be influenced. Repeated application of fertilizer can induce ion

exchange reactions, complexation processes and influence arsenic uptake by crops.

3.6.2.9 Coastal Sand Dunes It is clear by reviewing the literature that naturally elevated arsenic in groundwater is

found under many variable aquifer geochemical and depositional conditions. Despite

the many occurrences of elevated arsenic groundwaters, the coastal sand dune

environment rarely receives much attention, but is of utmost importance at Stuarts Point

where they form part of the main water bearing zone. A study conducted by Tesoriero et

al. (2004) examines water quality in a similar geomorphic setting as Stuarts Point,

however it fails to address arsenic occurrence. Appleyard et al.’s (2004) study of arsenic

occurrence in coastal sand dunes in Perth is attributed to ASS oxidation. Thus, arsenic

distribution in this coastal environment is controlled by a dominant and identifiable



geochemical process, differing from the variable conditions observed in the Stuarts

Point sandy aquifer which does not contain large amounts of distinct ASS horizons

(although they are present in close proximity to the sandy aquifer). The seemingly

homogenous sands derived from sediment movement onshore to form the Stuarts Point

aquifer is a unique and potentially unidentified environment for naturally high-As

groundwaters. A thorough search of scientific research databases provided no studies on

arsenic occurrence from any other coastal sand dune environment.

3.6.3 Arsenic in Australia There have been few studies on natural arsenic occurrences in the Australian

environment. Smith et al. (2003a) provide an overview of arsenic in Australia, stating

that,

“By far the most common source of elevated As concentrations in the Australian

environment are attributable to anthropogenic activities.” (pg 235)

Limited groundwater studies have been conducted and generally target past mining

activities. It appears that the Stuarts Point aquifer may be the first Australian

groundwater environment to receive research attention due to natural arsenic

occurrence. Subsequently, members of the same research group, the UNSW

Groundwater Group, have since noted elevated arsenic elsewhere in Australia. McLean

and Jankowski (2001) studied an inland alluvial aquifer used extensively for crop

irrigation in Australia and found it to contain approximately 70 g L-1 dissolved arsenic

from an unknown but presumably natural source. Morgan (2005) also noted elevated

arsenic concentrations in deep saline groundwaters responsible for producing extensive

dryland salinity in western NSW. Additionally, Groves (in prep.) and Gosavi (2004)

note arsenic is ubiquitous in ASS environments on the southeast Queensland coast, a

point also confirmed in coastal Perth by Appleyard (2004).

3.7 ARSENIC GEOCHEMICAL SUMMARY

Both Australian and coastal sand environments are little investigated in relation to

arsenic occurrence in groundwater and drinking water supplies. This literature review

has provided background information on the ubiquitous occurrence of arsenic in the



environment and the complexities involved in determining arsenic geochemical cycles.

Potential sources of arsenic in the Stuarts Point aquifer may include anthropogenic

contamination, sorption to aquifer materials during sea level transgressions and/or

erosion and deposition of regional mineral deposits. Sorbent materials contained within

the aquifer include clays, oxyhydroxides, calcite; and potential sinks like mineral

precipitates and arsenian pyrite. The mobilisation of arsenic from these sinks will vary

according to solution pH, Eh and composition. This chapter, summarising theoretical

information in conjunction with recent arsenic geochemical case studies from current

literature sources, provides sufficient background information required to make a

reliable assessment of arsenic source, retardation and mobilisation within the Stuarts

Point aquifer.


4 METHODOLOGY

4.1 CONSTRUCTION & INSTALLATION OF SAMPLING NETWORK

Ten multi-level piezometers were installed for this study and named ML1 through

ML10 (Figure 2.1). Sample locations for ML1-ML6 were selected to target seawater

intrusion, whereas ML7-ML10 sample locations were based on analysis of arsenic “hot

spots” delineated by Smith et al. (2003). ML1-ML6 were installed by the New South

Wales Department of Land and Water Conservation (now DIPNR) and correspond to

bore numbers GW081067 - GW081072 as delineated by that Department. ML7-ML10

were installed by the author herein. The piezometer design and installation methods

described below pertain to the latter multi-levels, however, those installed by DIPNR

followed similar procedures.

4.1.1 Multi-Level Piezometer Design

The multi-level piezometers consist of a 50 mm PVC pipe surrounded by 8 mm flexible

tubing extending to depth increments of 1 m and taped to the outside of the PVC casing

(Figure 4.1). The bottom 0.5 m of PVC is machine slotted and covered with a filter

sock. The ends of the 8 mm flexible tubing were hole punched, inserted with a scour

filter and covered in gauze bandage which acts as an additional filter. Photographs of

the design are provided in Appendix C.

4.1.2 Drilling

Drilling contractor’s (McDermott’s Drilling) were engaged to drill to depths of

approximately 30 m or until bedrock was penetrated. Table 4.1 shows the depth reached

for each multi-level and its respective sampling intervals. Drilling for ML7-10 was

undertaken by hollow flight augers without the use of any drilling fluid. Water was

required to stabilise the borehole after penetrating depths of approximately 10 m.

Drilling took place in July 2001. Piezometers were left to equilibrate for at least 6

months before sampling began. ML1-6 were installed the following year by mud rotary

method and supervised by hydrogeologists from DIPNR.

Chapter 4 - Methodology


Figure 4.1 Schematic of the multi-level design used at Stuarts Point.



Table 4.1 Depth and screened intervals for the multi-levels used in this study. Note that a screened interval of 5-30 indicates that a groundwater sample

was collected at metre increments between 5 m depth and 30 m depth, unless otherwise noted.

Multi-Level Depth (m) Screened Intervals (m)

ML1 32 5-30 *

ML2 31 5-27

ML3 31 3-27

ML4 34 3-30

ML5 33 4-30

ML6 40 7-30

ML7 27 6-28

ML8 30 6-29

ML9 24 5.5-21.5 ^

ML10 31 5-23 #

* Sample points 22 and 23 did not draw water ^ Sample point 11.5 did not draw water

# Sample points 8, 14-15 and 17-20 did not draw water

4.2 SEDIMENT AND GROUNDWATER SAMPLING

4.2.1 Sediment Sampling, Storage and Preparation Methods

During drilling, 54 sediment samples from ML7-ML10 were collected every 1.5 m via

split spoon sampling methods. Sediment pH was measured with a soil pH probe and

each sample was appropriately logged to compile the borelogs used throughout this

thesis. Samples were appropriately labelled in plastic bags, stored on ice and kept in the

dark until they could be frozen upon return to the laboratory.

4.2.2 Groundwater Sampling, Preservation and Storage Methods

In May 2001, 88 regional groundwater wells were measured with a dip meter to

determine groundwater levels. These were corrected to Australian Height Datum and

flow directions deduced from natural hydraulic gradients.

After the newly installed piezometers had equilibrated with the surrounding aquifer,

each of the ten multi-levels were sampled (n=226). Groundwater was extracted with a



Peristaltic GeoPumpTM attached to each 8 mm sample tube. Each sample point was

purged to remove standing water and allow groundwater representative of the aquifer to

be sampled. Sampling began after general parameters (temperature, pH and EC) had

stablised to within +/-5% of the previously measured reading. Each sample was field-

filtered to enable measurement of ‘dissolved’ species; dissolved being defined as those

substances that will pass through a 0.45 m MilliporeTM cellulose acetate membrane

filter. However, it is noted that colloids can range from 0.01 to 10 m in diameter

(Sposito, 1989), with Gschwend and Reynolds (1987) reporting colloid particles of

intermediate diameter – 0.1 m to 1.0 m – as being the most mobile particles in a

sandy medium. This was taken into account during the interpretation procedure.

Each sample was placed into acid washed plastic containers after rinsing first with the

actual groundwater to be sampled. Care was taken to minimise exposure of the sample

to the atmosphere, which may induce chemical changes prior to sample preservation.

Concentrated analytical grade nitric acid (10 drops of 32M nitric acid) was used to

preserve samples to eliminate precipitation and adsorption occurring within the matrix.

All groundwater samples were labelled accordingly and kept in chilled, dark, eskies

until transferred to refrigerators. Routine duplicates and blanks were submitted with the

samples to assess quality control procedures.

4.3 AQUEOUS CHEMICAL ANALYSES

4.3.1 General Parameters

Eight Orion meters (two for each parameter) were used to measure Eh, EC, temperature,

pH and DO. The average of the two readings is reported within. Probes were kept in de-

ionised water between samples and thoroughly washed to prevent cross-contamination.

Each meter was calibrated before going into the field and re-calibrated if the reported

measurements were more than +/-5% of the other meter. EC readings were measured

with a Model 122 conductivity cell and calibrated by checking measurements against

freshly prepared KCl standards ranging from 0.0005 to 0.2M. This follows the

calibration method for conductivity meters provided in APHA (1992). DO was

measured with a Model 820 oxygen meter, while pH and Eh were measured with a



Model 290A portable pH/concentration meter. pH meters were calibrated every day

using standard buffer solutions of pH 4.01, 7.00 and 10.01. A platinum redox electrode

was used to measure Eh after being calibrated against standard Zobell’s solution

(Nordstrom, 1997; Zobell, 1946).

4.3.2 Unstable Chemical Species

Redox sensitive parameters were measured at the time of sampling with a field-portable

Hach Spectrophotometer. Ferrous iron was measured using the 1,10 Phenanthroline

Method (Hach 8146) adapted from APHA (1992); sulfide was measured by the

Methylene Blue method (Hach 8131) adapted from APHA (1992); ammonium was

measured by the Nessler method (Hach 8038) adapted from APHA (1992); nitrate was

measured by the Cadmium Reduction method (Hach 8039); and phosphate the Reactive

Phosphorous method (Hach 8048) adapted from APHA (1992).

Additionally, field titrations were carried out at the piezometer head. The carbonate

species (bicarbonate and/or carbonate) were determined by field titration using 0.01M

or 0.1M HCl against methyl orange and bromocresol green indicators (APHA, 1992).

Carbon dioxide was measured in a similar manner against 0.025M NaOH with

phenolphthalein indictor (APHA, 1992). Each titration was performed twice and the

average of the two results was taken. All equipment was washed in nitric acid and

rinsed with both de-ionised water and sample water prior to use.

4.3.3 Major Ions

Major cations (Na+, K+, Ca2+ and Mg2+) were analysed by ICP-AES in the School of

Biological, Earth and Environmental Sciences at the University of New South Wales.

The machine was calibrated for the range of EC’s encountered and repeated

measurements were preformed on both samples and blanks (one blank for every ten

samples to account for machine drift) to monitor integrity of the analytical results.

Argentometric determination of chloride was ascertained by tritrating 0.01M AgNO3

against 10% potassium chromate indicator (APHA, 1992).



An overall check of precision and accuracy of the chemical analyses was conducted by

calculating charge balance errors (CBE’s) for the major ions to assess the

electroneutrality of each solution. CBE’s were calculated according to the following

equation:

CBE % = ( cations – anions) x 100

( cations + anions)

CBE’s not within +/-5% (Freeze and Cherry, 1979) were re-analysed for all major ions.

4.3.4 Minor Elements

Metals and trace elements (including arsenic) were analysed by ICP-MS in the School

of Chemical Sciences at the University of New South Wales. The same calibration and

quality control procedures described for the ICP-AES were applied by the laboratory

technician during the ICP-MS analyses. A commonly known problem involved in

analysing arsenic by ICP-MS is potential interference from chloride. The argon plasma

may form argon chloride, which has the same atomic mass (75) as arsenic. This

interference can raise reported arsenic concentrations by 1 g L-1 for every 100 mg L-1

of chloride present. Given the relatively fresh and low chloride concentrations of most

groundwaters at Stuarts Point this was initially not deemed a problem. However,

additional sampling of the seawater intrusion groundwaters and estuarine waters meant

this interference needed to be addressed. Corrections may be applied using the chlorine

isotope ratio but arsenic concentrations may remain inaccurate at the g L-1 level.

Therefore, the problem was addressed by calibrating the ICP-MS against a ‘seawater’

solution containing 19,000 mg L-1 Cl-. This ‘seawater’ solution was subsequently spiked

with arsenic concentrations ranging from 1 g L-1 - 1,000 g L-1 As (also selenium and

chromium as an additional check), diluted 100 times, and used to calibrate the machine

before analysis. For comparison, saline groundwaters from the adjacent wetland

previously analysed on the same ICP-MS and by the same technician reported no

concentrations of arsenic, thereby dismissing the increased arsenic bias in saline water

samples.



4.3.5 Arsenic Speciation

The original arsenic speciation method employed was also utilised by Smith et al.

(2003). The author herein conducted the analyses for both studies, but later found them

to be questionable due to chemical instability and concern over the accuracy of the

original method - APHA Method 3500-As B (20th ed., 1998). Custom made glassware

enabled a larger groundwater sample to be analysed and ensured equipment was robust

to sustain field conditions. After preparing spiked arsenic solutions in the laboratory, a

calibration curve was constructed which enabled the field-measured arsenic

concentrations to be determined. This laboratory work was initally conducted at the

Australian Nuclear Science and Technology Organisation under Dr Maree Emmett. The

following list is the step-by-step procedure carried out during As(III) analysis in the

field at Stuarts Point:

1. Place 400 mL of sample in glass jar and place onto magnetic stirrer.

2. Grease reaction vessel, place all stoppers in holes, add custom made absorber tube.

3. Add 5 mL of acetate buffer to sample by pipette.

4. Place 10 g of 3 mm glass beads into absorber tube (making sure lead acetate wool is

in the mouth of the absorber tube).

5. Record absorbance of DTC solution (as a blank) at 520 nm in the

spectrophotometer.

6. Add 20 mL of DTC solution from a glass syringe into absorber tube.

7. Connect and adjust nitrogen gas flow into reaction vessel to 60 mL per minute,

making sure gas is bubbling through DTC solution.

8. Fill a 60 mL plastic syringe with 30 mL of sodium borohydride and connect to the

nitrogen gas line.

9. Start stopwatch.

10. Slowly (over 2-5 minutes) add the sodium borohydride to the groundwater sample.

11. Leave to develop over 20 minutes.

12. At 20 minutes, redissolve any crystals in the absorber tube by pulling one of the

glass stoppers in and out, releasing pressure in the glass cylinder.

13. Place DTC solution into spectrophotometer cell and measure the absorbance at

520nm.

14. Determine As(III) concentration off previously prepared absorbance curve.



4.4 SOLID PHASE ANALYSES

4.4.1 Electron Microscopy

All electron microscope work was undertaken by the author (with the assistance of Mr

Barry Searle) at the Electron Microscope Unit (EMU) within the University of New

South Wales (UNSW). The following section details sample preparation procedures and

instrument particulars, and is largely drawn from the in-house manual compiled for

EMU users at UNSW (Munroe and Stevens-Kalceff, 2002).

4.4.1.1 Sample Preparation

Fifteen samples were analysed via Scanning Electron Microscope – Energy Dispersive

x-ray Spectrometer (SEM-EDS). Each sample was mounted on an aluminium stub 2.5

cm in diameter, by adherence to double-sided carbon tape. The samples were then

placed in a sputter coater and coated with a thin layer of carbon to make them

electrically conductive. Carbon was chosen as the coating element as other metallic

coats (such as Au, Pt or Cr) may interfere with the chemical analysis.

For samples necessitating more accurate chemical analysis (n=10), the electron

microprobe equipped with a Wavelength Dispersive x-ray Spectrometer (WDS)

required samples to be embedded in a resin for polishing. A cross sectional sample was

thus prepared by Mr Rad Flossman in the School of Biological, Earth and

Environmental Sciences at UNSW. These samples were processed so as to be optically

flat, before coating with carbon to provide a path for the probe current to flow to the

earth.

4.4.1.2 SEM-EDS

SEM analysis was conducted on a Hitachi S4500 Field Emission SEM (1996), with

high resolution (1.5 nanometers), a tilting stage, Robinson Back-Scatter Detector,

Oxford Cathodoluminescence Detector (MonoCL2/ISIS) and a Link ISIS 200

Microanalysis System for chemical determination. The accelerating voltage was set at

20kV throughout the analysis process. Spectra were analysed by comparing the energies

of the observed peaks with characteristic peak energies of the elements. Care was taken



with elements that had overlapping peaks, such as arsenic and lead; several peaks were

examined in order to identify the correct element spectra produced.

4.4.1.3 Electron Microprobe-WDS

The electron microprobe consists of a SEM with several WDS interfaces, giving it the

advantage of a more accurate and sensitive chemical analysis than the standard SEM-

EDS analysis. However, the WDS detector can only detect one element at a time,

whereas the EDS can semi-quantitate several elements at a time. Thus, the WDS is a

more powerful tool but is both costly and timely in its analysis.

A Cameca SX50 electron microprobe employing four WDS’s and one EDS was utilised

to produce a line scan over a grain with an identified coating layer. Several individual

points were also analysed quantitatively. Characteristic x-rays generated by elements

within the sample are diffracted by a suitable analysing crystal and measured in counts

per second by the gas flow proportional detectors.

4.4.1.4 Internal QA/QC Procedures

Both instruments are calibrated with Standard Reference Materials for the various

elements analysed. The S4500 uses Guide E1508-98 Standard Guide for Quantitative

Analysis by Energy-Dispersive Spectroscopy; while the SX50 employs the data

reduction matrix correction procedure based on the methods of Pouchou and Pichoir

(1985).

Various steps in the SX50 calibration process were carried out by the technical officer

in charge of the instrument (Mr Barry Searle). This required manual entry of calibration

criteria such as background positions, counting time, background slope and the

analytical line. The acquisition parameters used during calibration (including crystals

used and interferences accounted for) may be supplied on request.

Detection limits for the SX50 may be affected by random errors such as calibration

uncertainty, peak and background counting time, system stability, profile slope,

uncertainty in any overlap corrections, the accelerating voltage, the beam current, the



line used to measure the element and the composition of both the sample and the

standards. The peak counts must exceed three times the standard deviation of the

background to give a 95% probability that a peak may be present. Only those elements

adhering to this criteria were given a reported concentration value; all others were

reported as 0%. The detection limits for the SX50 are generally set at 0.01% (100ppm)

except for As which was set at 0.02%. This compares with the S4500 which has

detection limits of 0.2%.

4.4.2 Grain Size Analysis

Nineteen samples were selected for grain size analysis (GSA). The sand and granule

fractions were targeted during the GSA procedure in order to differentiate beach sands

from river sands. Approximately 50-100 g of split sample was first wet-sieved through a

63 m (0.063 mm) sized sieve to remove the silt and clay fractions. These two fractions

are therefore reported together as a percentage of each respective sample. The

remaining sand and granule fractions were oven dried and sieved through a set of 8-inch

diameter sieves at ½ phi ( ) intervals and placed on an electronic shaker for 5 minutes.

Each respective size fraction was weighed to 0.01 g.

Histograms and cumulative frequency curves were generated for the grain size results.

An arithmetic scale was used for cumulative frequency. The parameters of mean grain

size, skewness and standard deviation (sorting) were determined using the graphical

formulae of Folk and Ward (1957) with grain size measured in . As an alternative, they

were also calculated using the method of moments (as cited in Tucker, 1991). Formulae

for both methods are presented in Table 4.2. A comparison between the results of each

method are shown by the mean relative percent difference (RPD) calculated for each

parameter (Table 4.3). These calculations are provided in Appendix B3.



Table 4.2 Calculation of various grain size parameters using different methods. Grain Size Parameter

Graphical Method Folk and Ward (1957)

Method of Moments (as cited in Tucker, 1991)

Mean ( ) = 16 + 50 + 84

3

= fm

100

Standard

Deviation ( )

= 84 - 16 + 95 - 5

4 6.6

= f(m - )2

100

Skewness ( 3) 3 = 16+ 84 - 2 50 + 5+ 95 - 2 50

2( 84 - 16) 2( 95 - 5)

3 = f(m - )3

100 3

Table 4.3 Mean RPD values between calculations by the graphical and moment measures methods.

Parameter RPD %

Mean ( ) 4.34

Standard Deviation ( ) 0.23

Skewness ( 3) 0.03

Mean RPD’s are below 5% indicating the relative similarities between each

mathematical method. The method of moments was chosen for all statistical

comparisons via bivariate plotting techniques. Folk (1966) makes reference to some

disadvantages with the method of moments; namely errors associated with sieve screen

sizes and problems encountered when sediment is finer than 4 . The latter problem has

been removed herein when differentiating between beach and river sands by using only

those samples that have grain sizes coarser than 4 (i.e., the sandy samples only) and

hence omitting the clayey estuarine samples.

In addition, studies by Hails (1967) found that the results of both the graphical and

moment measures successfully differentiated between Holocene and Pleistocene

sediments, with only minor variation occurring between mean grain sizes and sorting



values. Thus the selection of one mathematical technique over the other is insignificant

to the final result in this instance.

4.4.3 Sequential Extractions

A modified (cation/anion) sequential extraction technique (Table 4.4) was applied to

specifically target arsenic pools in the sediments. This procedure was chosen based on a

literature review of sequential extraction schemes.

Table 4.1 lists references cited and common advantages/disadvantages for each targeted

fraction. Due to the disputable nature of such extraction techniques, only five samples

were chosen for this procedure. Each sample was significantly different to the other

(brown sand, green sand and clay, orange/red clay, coffee rock and farm topsoil), and

were chosen for analysis as a preliminary screening tool leading to further analytical

techniques.

4.4.4 XRD

XRD analysis was carried out on a Philips PW1140 Diffractometer located in the

School of Biological, Earth and Environmental Sciences at the University of New South

Wales. Sediments were first oven-dried and ground to pass a 2 mm sieve. Semi-

quantitative mineral analysis was aided with XPLOT software providing a database of

representative XRD mineral peaks that were visually matched to the resulting analytical

spectra.

4.4.5 XRF

Both major and minor elements were analysed in samples on a Philips PW2400 XRF

with a 102 position sample changer by Ms Irene Wainwright in the School of

Biological, Earth and Environmental Sciences at the University of New South Wales.

Sample preparation involved mechanical grinding of sediments prior to construction of

‘pellets’ for trace element analysis and ‘glass plates’ for major elements.



Table 4.4 Step-by-step sequential extraction scheme utilised herein.Fraction Method

SamplePreparation

Remove soils from freezer and air dry Grind/crush soils, shells and rock fragments before passing through a 2 mm sieve Weigh 1 g of soil into 50 mL polypropylene centrifuge tube with sealable lid

Exchangeable Fraction

Pipette 25 mL of (NH4)H2PO4 into the centrifuge tube and place tube on shaker for 16 hours

After shaking centrifuge at 3000 rpm for 15 min Filter the supernatant through a 0.45 m filter into a PE-bottle Acidify with HNO3 ready for ICP-MS analysis – store samples in dark fridge until

analysis Wash the residue twice with 10 mL of Milli-Q water and add washing liquid to the

already extracted solution

Carbonate Fraction

Using the residue from the previous extraction, pipette 25 mL of 1M sodium acetate (CH3COONa – adjusted to pH 5 with acetic acid) into the tube and shake for 5 hours

After shaking centrifuge at 3000 rpm for 15 min Filter the supernatant through a 0.45 m filter into a PE-bottle Acidify with HNO3 ready for ICP-MS analysis – store samples in dark fridge until



Reducible Fraction

Using the residue from the previous extraction, pipette 30 mL of 0.5M sodium citrate and 2.5 mL of NaHCO3 into the tube while adding 0.5 g of sodium dithionite

Heat in a sand/water bath at 85 C for 15 minutes Centrifuge at 3000 rpm for 15 min Filter the supernatant through a 0.45 m filter into a PE-bottle Acidify with HNO3 ready for ICP-MS analysis – store samples in dark fridge until



Oxidisable Fraction

Using the residue from the previous extraction, add 3 mL of 0.02M HNO3 and 5mL of 30% H2O2, adjusting the pH to 2 with HNO3

Heat in a sand/water bath at 85 C for 2 hours and occasionally agitate After 2 hours add 3 mL of 30% H2O2 (pH adjusted to 2 with HNO3) and continue

heating for another 3 hours, agitating occasionally After 3 hours (total of 5 hours heating) let the sample cool, then add 5 mL of 3.2M

ammonium acetate in 20% (v/v) HNO3 and dilute to 20 mL with Milli-Q water Shake for 30 minutes at room temperature Centrifuge at 3000 rpm for 15 min Filter the supernatant through a 0.45 m filter into a PE-bottle Acidify with HNO3 ready for ICP-MS analysis – store samples in dark fridge until



Residual Fraction

Determine total elements by XRF and subtract fractions 1-5 from XRF total to give the residual fraction

Tab

le 4

.5 C

omm

only

app

lied

sequ

entia

l ext

ract

ion

tech

niqu

es a

nd th

eir

asso

ciat

ed a

dvan

tage

s/di

sadv

anta

ges.

Ta

rget

edFr

actio

nC

omm

only

Use

d Ex

trac

tion

Che

mic

als

Adv

anta

ges

/ Dis

adva

ntag

es

Cite

d In

Exc

hang

eabl

eFr

actio

nA

mm

oniu

m o

r sod

ium

ace

tate

Mag

nesi

um c

hlor

ide

May

als

o at

tack

car

bona

tes

Ace

tate

com

plex

es w

ith c

alci

um

Chl

orid

e m

ay in

terfe

re w

ith IC

P-M

S

Tess

ier e

t al.

(197

9); C

ai e

t al.

(200

2)

Tess

ier e

t al.

(197

9)

Car

bona

teFr

actio

nS

odiu

m a

ceta

te o

r ace

tic a

cid

Low

er p

H v

alue

s pa

rtial

ly a

ttack

Fe

& M

n ox

ides

Doe

s no

t affe

ct s

ilica

tes,

sul

fides

or o

rgan

ic m

atte

r

Tess

ier e

t al.

(197

9); L

a Fo

rce

et a

l. (2

000)

Wen

zel (

2001

)

Red

ucib

leFr

actio

nS

odiu

m d

ithio

nite

citr

ate

Hyd

roxy

lam

ine

hydr

ochl

orid

e

Sub

stan

tial p

reci

pita

tion

of tr

ace

met

als

with

inso

lubl

e su

lfide

s fro

m d

ispr

opor

tiona

tion

of d

ithio

nite

, whi

ch is

als

o hi

ghly

con

tam

inat

ed w

ith z

inc

Chl

orid

e m

ay in

terfe

re w

ith IC

P-M

S

Tess

ier e

t al.

(197

9); G

leyz

es e

t al

. (20

02a)

Tess

ier e

t al.

(197

9)

Oxi

disa

ble

Frac

tion

Hyd

roge

n pe

roxi

de in

aci

dic

med

ium

Con

cent

rate

d ni

tric

or

hydr

ochl

oric

or p

erch

loric

aci

d

May

not

oxi

dise

all

form

s of

org

anic

mat

ter,

also

ext

ract

s so

me

sulfi

de m

iner

als

Als

o at

tack

s so

me

silic

ate

latti

ces

in a

dditi

on to

pot

entia

l ch

lorid

e in

terfe

renc

e w

ith IC

P-M

S

Tess

ier e

t al.

(197

9); G

leyz

es e

t al

. (20

02a)

Tess

ier e

t al.

(197

9)

Res

idua

lFr

actio

nH

F or

nitr

ic a

cid

or p

erch

loric

ac

id

XR

F

HF

high

ly to

xic

and

pote

ntia

l chl

orid

e in

terfe

renc

e

Muc

h sa

fer t

han

usin

g H

F

Tess

ier e

t al.

(197

9)

Stu

mm

eyer

et a

l. (2

002)



4.4.6 Loss On Ignition (LOI)

LOI was determined as part of the XRF analytical process and can be used as an

indication of the amount of organic matter in the sample - roughly 40-60% LOI may be

organic carbon (Bengtsson and Enell, 1986). Such a method has been used by Welch

and Lico (1998) to determine arsenic and uranium occurrence in shallow groundwater.

However, given the amount of carbonates, sulfides and potential hydration minerals in

these sediments, the use of LOI as an indicator of organic matter content is not

considered plausible here.

4.5 STATISTICAL ANALYSES

The following statistical methods were applied by the author herein in a separate

hydrogeochemical study (O’Shea and Jankowski, 2006).

4.5.1 Data Screening

4.5.1.1 Assessment Of Normality

Prior to commencing statistical analyses, the data set should be assessed for normality

and completeness of data. Results of the data screening process determines if the data is

suitable for statistical tests in raw form, or if it requires manipulation to achieve more

accurate results from the statistical procedures employed. The data screening process for

the Stuarts Point chemical data is described below. Major ions and trace elements were

assessed independently due to large compositional differences.

Data producing a bell shaped frequency distribution, with values clustered around a

central point and the frequency of occurrence declining away from this point, are said to

be normally distributed (Davis, 1986). It is often assumed that variables are normally

distributed, and many statistical tests are based on this assumption. PCA is an exception

to this rule as it is based solely on eigenanalysis of the correlation/covariance matrix

(Meglin, 1991). However, HCA assumes the data is either normal or log-normal (Guler

et al., 2002), therefore an assessment of normality was required before HCA could be

undertaken.



Histograms produced in SPSS showed that the major ion groundwater data from Stuarts

Point are positively skewed. Most naturally occurring element distributions follow this

pattern (Miesch, 1976). An assessment of normality for the major ions indicated that

neither the raw nor log10-transformed data appeared to approximate normality. Based on

these results, the major ion groundwater data was left in raw form for the statistical

analyses. In contrast, the trace element groundwater data showed a closer approximation

to normality after log10-transformation. There are advantages and disadvantages to

transformating data sets. Dreher (2003) debated the use of transformation techniques on

positively skewed data due to the possibility of information – specifically anomalies –

to become lost. Guler and Thyne (2003) responded to Dreher’s comment by stating that

transformation is required to produce maximum results from parametric (normal)

statistical methods. Given debate in this subject, the raw trace element data was used for

PCA (which does not require the assumption of normality) and the log10-transformed

data was used for HCA (which does require the assumption of normality). When both

data sets were combined transformed data was used; as was the case for the sediment

samples.

4.5.1.2 Standardization

In order for each variable to have an equal weight in the statistical analyses, the data can

be standardized to a range of –3 to +3 standard deviations, centred about a mean of zero

(Guler et al., 2002). Dreher (2003) disputed the use of standardization, saying it shifts

the relationships between the ions and neglects information derived from relative

abundances, such as supply and demand on the air, soil, unsaturated and saturated

zones. Guler and Thyne (2003) responded to this comment saying they used

standardization to avoid misclassifications arising from the inappropriate weighting of

parameters with large magnitudes. A number of other authors have standardized their

geochemical data (Farnham et al., 2000; Ruiz et al., 1990; Swanson et al., 2001).

Standardization is necessary when data varies over wide ranges, as is the case with

concentrations of major ions and trace elements in both the Stuarts Point groundwater

and sediment data sets. Standardization was employed herein to ensure each analyte was

equally weighted during the statistical analyses.



4.5.1.3 Censored Data

Robust statistical techniques achieve the most reliable results when a full (i.e.,

complete) data set is used. However, as is often the case in geochemical investigations,

the Stuarts Point data set contains ‘censored’ values. Values reported as “less than” or

“greater than” a censor point are known as censored or qualified values (Sanford et al.,

1993). The censor point for these investigations is the analytical detection limit (DL) for

each variable (analyte). Several analytes reported concentrations at or below the

analytical DL and were thus assigned a value of “less-than” the DL. Such censored data

can complicate all subsequent statistical analyses (Farnham et al., 2002) and therefore

consideration should be given to replacing the censored data with uncensored values.

Sanford et al. (1993) showed the importance of treating censored data as opposed to

ignoring it. They discuss the importance of replacing censored values, to prevent

misguided statistical results.

A number of techniques are available for replacing censored data values, depending on

the size of the data set and the number of censored values present. Perhaps the most

common method is replacement of the less-than values with ¾ times the lower DL and

the greater-than values with 4/3 times the upper DL (VanTrump and Miesch, 1977).

Similar to this, replacement with half the DL is the current method supported by the

New South Wales Environmental Protection Authority (NSW EPA, 1995) for the

assessment of contaminated sites. Additionally, replacement with zero or the DL are

also common substitution values. Seyhan et al. (1985) settled on substitution with the

DL stating that this has an insignificant effect on the Q-mode classification results of

principle components, cluster and discriminant analysis. However, Farnham et al.

(2002) showed that substitution with half the detection limit (DL/2) gave superior

results when compared to the DL or zero. Significant deterioration of the performance

of all substitution methods occurred when the number of less-than values exceeded

25%. Sanford et al. (1993) proposed that for a small proportion of substitutions, a

replacement factor of 0.55 is preferable to using ¾ the DL.

An alternative approach is to use the actual reported concentration produced by the

analytical instrument, even if it is below the DL. Many chemists won’t report these less-



than values stating they are unreliable. However, studies have shown that these

‘unreliable’ values can more effectively detect trends than censored data (Gilliom et al.,

1984). These values were used for the Stuarts Point sediment concentrations reported by

the XRF.

More precise methods are available for determining descriptive statistical values (such

as mean and variance) but are not commonly used for multivariate techniques. These

methods involve estimating the mean of the (normally distributed) censored data set and

using this estimated mean as a replacement value. Further information on these

statistical procedures can be found in Farnham et al. (2002), Sanford et al. (1993) and

Singh and Nocerino (2002).

An assessment of the raw major ion groundwater data showed three values for

potassium were reported as being below the analytical DL (<0.16 mg L-1). Given the

low frequency of replacement required (less than 2% of the data), a replacement value

of half the detection limit (DL/2 - 0.08 mg L-1) was used for the three censored

potassium values. This method follows that used by Farnham et al. (2002) and the NSW

EPA (1995). This value was also extremely close to using a replacement factor of 0.55

as recommended by Sanford et al. (1993), which would yield a value of 0.088 mg L-1.

Replacement with DL/2 was also selected for the trace element data set. This method

was selected based on Farnham et al.’s (2002) assessment of the performance and

deterioration of replacement values. Given the deterioration of substitution methods

when increasing amounts of censored values are present, elements with more than 30%

replacements required were removed from the statistical analyses. The same approach

was taken by Critto et al. (2003) who removed those parameters showing low frequency

of detection from their statistical analysis.

4.5.2 Descriptive Statistics

Prior to subjecting the Stuarts Point data to multivariate statistical manipulation, an

analysis of descriptive statistics was conducted. Descriptive statistics provide an

indication of abundance and elemental outliers. All statistical analyses were performed

using the software package SPSS Version 12.01.



4.5.3 Multi-Variate Statistical Analyses

4.5.3.1 Cluster Analysis

Hierarchial Cluster Analysis was performed on all sediment and groundwater samples.

Q-mode HCA classified each case (sample), measured by variable (element or ion) into

statistically defined groups. Ward’s (1963) linkage method was chosen to iteratively

link nearby points of similarity. Euclidean distance measure assessed the similarity

combined with linkage. R-mode HCA was useful in establishing relationships between

the variables (analytes) rather than the cases (samples) as is the situation in Q-mode

HCA. A dendrogram was produced for each cluster analysis undertaken. In most cases,

the phenon line was set at an Euclidean distance of five to identify clusters of most

similarity.

4.5.3.2 Principal Components Analysis (PCA)

A second geostatistical tool was used on the Stuarts Point data – Principal Components

Analysis. PCA is a mathematical technique used for reducing data and deciphering

patterns within large data sets (Hull, 1984; Joliffe, 1986; Stetzenbach et al., 1999; and

Wold et al., 1987). Principal Components (PC’s) are based on eigenanalysis of the

correlation or covariance matrix, therefore data does not need to be normally distributed

(Meglin, 1991). PCA was thus conducted on raw standardized data using the

correlation matrix. PC’s are the result of strong correlations between variables. PC’s

may result from the correlation of suites of variables (such as marine elements)

representing the same geological origin or geochemical source. The first PC is that

component which has the greatest possible variance, the second PC has the second

greatest variance, and so on. All PC’s are uncorrelated (ie, orthogonal) to one another

(Stetzenbach et al., 2001). Eigenvalues describe the amount of variance explained by

each PC, and thus decrease with each successive PC extracted. Eigenvectors (or PC

loadings) indicate the relative contribution each element makes to that PC score. The

larger the score - in absolute value - the stronger the influence of that element (Webster,

2001). While PCA was utilised as a pre-screening device for the hydrochemical

interpretation, the best statistical results were provided by HCA and thus the results of

the PCA are not included within.



4.5.4 Statistical Correlation

Most geochemical data are positively skewed. As such, Spearmans correlations are

often used in geochemical correlation. However, Pearsons correlations are better suited

to normal data. Since the Stuarts Point data has been log10-transformed to better

approximate normality, Pearsons correlation technique is most appropriate. Pearsons

correlations were calculated in SPSS. All r2 values provided on graphs, tables and in the

text, are correlations performed in SPSS, not Excel.


5 ORIGINAL SOURCE OF ARSENIC TO THE AQUIFER

5.1 HYPOTHESES

Five arsenic source theories are hypothesised:

1. Arsenic is present in the aquifer as a by-product of anthropogenic activities;

2. Arsenic has been contributed to the aquifer matrix via deposition of regionally

eroded geological units containing arsenic mineralisation;

3. Arsenic is derived from remnant seawater trapped in marine clay units deposited

during eustatic changes of sea level in the Quaternary (Smith et al., 2006);

4. The oxidation of arsenian pyrite present in ASS material contributes dissolved

arsenic to the groundwater (Smith et al., 2006); and/or

5. The underlying bedrock contains arsenic, which is being contributed to the aquifer

via upwards vertical leakage of groundwater (Smith et al., 2003).

This chapter discusses the validity of each hypothesis and determines which one is most

likely to be the dominant arsenic source in the aquifer. A number of investigative

techniques have been utilised herein in the assessment of each potential arsenic source.

Table 5.1 lists each proposed source, and the work completed herein, in order to assess

the validity of each source.

This chapter is structured as follows:

the results of investigative techniques are provided first;

followed by a discussion of each potential arsenic source and the evidence

collected both for, and against, each hypothesis;

the chapter concludes with a comparison to known arsenic sources in other

elevated arsenic environments; and finally

a summary of the findings presented in this chapter.

Tab

le 5

.1 T

echn

ique

s us

ed h

erei

n to

dis

mis

s or

sup

port

pot

entia

l ars

enic

sou

rces

to th

e St

uart

s Po

int a

quife

r, a

nd th

e lo

catio

n of

res

ults

pr

ovid

ed w

ithin

this

thes

is.

Pote

ntia

lA

rsen

ic S

ourc

e Sp

ecifi

c Q

uest

ion(

s) to

be

addr

esse

d he

rein

Inve

stig

ativ

e Te

chni

ques

Lo

catio

n of

Dat

a

Ant

hrop

ogen

ic

arse

nic

Has

ant

hrop

ogen

ic c

onta

min

atio

n oc

curr

ed?

Lite

ratu

re re

view

of g

over

nmen

t rec

ords

; ane

cdot

al

evid

ence

; pa

st m

inin

g hi

stor

y; in

terv

iew

s w

ith p

rese

nt

resi

dent

s

Ana

lysi

s of

gro

undw

ater

dat

a (s

uppo

rting

evi

denc

e)

Inte

grat

ed in

to d

iscu

ssio

n in

Sec

tion

5.3.

1

Gro

undw

ater

dat

a di

scus

sed

fully

in

Cha

pter

7

Reg

iona

l geo

logi

c so

urce

W

as th

e S

tuar

ts P

oint

aqu

ifer d

epos

ited

by o

nsho

re s

edim

ent m

ovem

ent,

as

prev

ious

ly p

ropo

sed,

or d

oes

it ha

ve

sign

ifica

nt fl

uvia

l inf

luen

ce?

If th

e aq

uife

r has

bee

n de

posi

ted

by

fluvi

al e

rosi

on o

f the

hin

terla

nd, i

s th

ere

regi

onal

ars

enic

min

eral

isat

ion

pres

ent?

Com

pila

tion

of b

orel

ogs

from

dril

ling

Sed

imen

t gra

in s

ize

anal

yses

to d

eter

min

e if

fluvi

al

sedi

men

tatio

n ha

s oc

curr

ed

Mol

lusc

iden

tific

atio

n to

sho

w s

peci

fic p

arts

of t

he

aqui

fer w

ere

depo

site

d un

der h

abita

t con

ditio

ns

suita

ble

for c

erta

in m

ollu

sc s

peci

es

Dev

elop

men

t of a

quife

r cro

ss s

ectio

ns

Det

erm

inin

g so

urce

pro

vena

nce

for t

he a

quife

r se

dim

ents

Dev

elop

men

t of a

n aq

uife

r spe

cific

geo

mor

phic

mod

el

Rev

iew

of r

egio

nal m

iner

alis

atio

n

Sec

tion

5.2.

1

Sec

tion

5.2.

2.1

Spe

cies

iden

tific

atio

n di

scus

sed

in

Sec

tion

5.2.

2.2

; loc

atio

ns o

f spe

cies

pr

ovid

ed o

n bo

relo

gs in

Sec

tion

5.2.

1

Sec

tion

5.2.

4.2

Sec

tion

5.2.

5

Sec

tion

5.2.

6

Sec

tion

5.2.

5.2

Pote

ntia

lA

rsen

ic S

ourc

e Sp

ecifi

c Q

uest

ion(

s) to

be

addr

esse

d he

rein

Inve

stig

ativ

e Te

chni

ques

Lo

catio

n of

Dat

a

Can

a li

nk b

e m

ade

betw

een

regi

onal

ar

seni

c an

d th

e pr

esen

ce o

f ars

enic

at

Stu

arts

Poi

nt?

Sta

tistic

al c

lust

erin

g of

sed

imen

t che

mis

try

Sel

ectio

n of

a h

inte

rland

indi

cato

r ele

men

t to

corr

elat

e aq

uife

r sed

imen

ts w

ith a

regi

onal

sou

rce

prov

enan

ce

Sed

imen

t che

mis

try re

sults

pro

vide

d in

Sec

tion

5.2.

3; s

tatis

tical

resu

lts

disc

usse

d in

Sec

tion

5.2.

4.1.

Sec

tion

5.2.

5.2

Rem

nant

se

awat

er

Did

sea

wat

er tr

ansg

ress

ove

r the

S

tuar

ts P

oint

aqu

ifer?

Do

the

curr

ent a

quife

r sed

imen

ts h

old

any

clue

s in

dica

ting

a pa

st o

rigin

from

se

awat

er?

Geo

mor

phic

reco

nstru

ctio

n of

the

Stu

arts

Poi

nt a

quife

r

Ana

lysi

s of

ars

enic

con

cent

ratio

n pe

aks

in s

edim

ents

an

d th

e di

strib

utio

n of

est

uarin

e cl

ay u

nits

Sec

tion

5.2.

6

Ars

enic

che

mis

try p

rovi

ded

in

Sec

tion

5.2.

3; c

orre

latio

n w

ith

bore

logs

(est

uarin

e cl

ay u

nits

) di

scus

sed

in S

ectio

n 5.

3.3

AS

S a

ssoc

iatio

n Is

ther

e an

y as

soci

atio

n w

ith A

SS

ho

rizon

s an

d ar

seni

c?

Is th

ere

any

asso

ciat

ion

with

pyr

ite a

nd

arse

nic?

Obs

erva

tion

of A

SS

hor

izon

s

SE

M u

sed

to id

entif

y py

rite

Bor

elog

s in

Sec

tion

5.2.

1; d

iscu

ssio

n in

Sec

tion

5.3.

4

Res

ults

pre

sent

ed in

Cha

pter

61

Dire

ct b

edro

ck

cont

ribut

ion

Is a

rsen

ic p

rese

nt in

the

bedr

ock?

Is a

rsen

ic b

eing

con

tribu

ted

to th

e un

cons

olid

ated

aqu

ifer v

ia d

isch

arge

fro

m th

e be

droc

k aq

uife

r bel

ow?

Rev

iew

of l

itera

ture

and

ana

lysi

s of

bed

rock

en

coun

tere

d du

ring

drill

ing

Inve

stig

atio

n of

sta

tistic

al c

orre

latio

n be

twee

n ar

seni

c an

d tra

ce e

lem

ents

in g

roun

dwat

er s

ampl

ed c

lose

to

bedr

ock

Sec

tion

5.2.

2.3

Ver

tical

gra

ph o

f an

indi

cato

r ele

men

t pr

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Chapter 5 – Original Source of Arsenic to the Aquifer


5.2 INVESTIGATIVE RESULTS AND INTERPRETATION

The aquifer matrix was determined to be the key medium to investigate both arsenic

source and current sinks (Chapter 6). The natural (geogenic) arsenic sources proposed

for the Stuarts Point aquifer are a direct result of the aquifer’s depositional history. To

date, some regional geologic and geomorphic surveys have been extended to encompass

Stuarts point (Roy and Thom, 1981; Hails, 1968), however, no aquifer-specific

geomorphic studies have been conducted. Data collected during this study allows the

first comprehensive geomorphic model to be developed for Stuarts Point.

5.2.1 Aquifer Lithology

To enable a complete geomorphic reconstruction, the installation of ten multi-level

piezometers ranging in depth from 24 – 40 metres below ground surface (mbgs)

allowed detailed logging of the aquifer sediments to be recorded. Six boreholes (ML1-

ML6) were installed and logged by the NSW DLWC to monitor seawater intrusion,

while the remaining four boreholes (ML7-ML10) were completed by the author

specifically for this study (see Chapter 4 for full drilling methodology). All ten

piezometers have been utilised in this investigation. The location of these multi-levels is

shown on the map provided in Chapter 2 (Figure 2.1).

Sediments were visually logged by collecting a sample every 1.5 m during drilling.

Individual borelogs are shown in Figure 5.1 through Figure 5.10.

Figure 5.1 Borelog for ML1.












Various sediment facies were observed during drilling:

An organic topsoil layer consisting of natural vegetative litter or domestic lawn

seed was present at all borehole locations;

Sand was penetrated below the organic soil layer. The sand ranged in colour

from dark brown (influenced by organic matter) to yellow, through to olive grey,

with increasing depth. The shallow brown and yellow sands were generally of

uniform sorting and fine to medium grained; quartz dominant. Some indurated

‘coffee rock’ was periodically observed where old water table fluctuations have

occurred, cementing organic matter with the sand grains. Sands olive grey in

colour and present at mid-depths in the aquifer often exhibited a fining upwards

grain size pattern and sometimes contained large lithic fragments; both angular

and round in shape.

Clay material was observed at mid to deeper depths in the aquifer. Sand units

became clayey and mollusc shells are often present. Rocks and pebbles are

sometimes interspersed with clay units. Highly plastic olive grey clay units are

encountered towards the base of the aquifer. Many boreholes are terminated in

weathered bedrock resembling phyllite or shale.

5.2.2 Aquifer Depositional Conditions

5.2.2.1 Differentiating between beach sand and fluvial sand

The sands encountered in the shallow section of the aquifer exhibited homogeneity in

both grain size and colour; typical of the barrier sand system of the NSW coastline

(Hails, 1967). Sediment beneath these proposed beach barrier sands changes into

coarser, poorly sorted sediments characteristic of fluvial depositional conditions. These

sediment differences are particularly noticeable in Figure 5.10. The first 8 m of sand are

uniformly sorted and composed mainly of quartz. Below 8 m, the sands become

increasingly poorly sorted, with large lithic fragments often present. These large

fragments are often indicative of flood deposits. By observing these changes in the field,

one can hypothesise that the shallow uniform sediments were deposited under a

different energy regime than the deeper non-uniform sediments.



However, a facies change observed in the field is not considered sufficient to

differentiate between fluvial and barrier depositional conditions. Some common grain

size parameters are often used by sedimentologists to aid in the differentiation of sands

from different environments, such as dune versus river sands. An example of one such

study is presented in Friedman (1961), who studied sands from recent dune, beach and

river environments with the aim of determining if certain mineralogical or textural

characteristics were diagnostic of particular depositional environments. His findings

(relevant to this study) include:

Dune and beach sands can almost completely be differentiated by a bivariate plot of

mean grain size and the third moment (skewness).

Both dune and river sands are generally positively skewed, however dune sands

tend to be better sorted than river sands.

Likewise, beach sands tend to be better sorted than river sands.

For all the sands studied by Friedman (1961), skewness reflects the environment of

deposition and is not affected by the mineralogy of the sample.

Bivariate plots of these grain parameters can thus indicate environments of deposition,

as utilised in studies by Hails (1967) and Smith (2005). The sands collected from the

aquifer were thus separated from clay samples and subjected to grain size analysis.

Grain parameters of skewness and standard deviation (sorting) were calculated by two

methods; the full methodology is described in Chapter 4. Results of these calculations

and a comparison between the two methods can be found in Appendix B3.

A plot of moment skewness versus moment standard deviation (Figure 5.11) identified

samples derived from beach processes (generally sands present at the surface ranging to

a depth of 10 mbgs) and those sands more likely to be a product of fluvial transport

mechanisms (sands from approximately 10 to 15 mbgs and another coarser layer

existing at around 20m depth). They have been termed ‘beach’ and ‘river’ sands on the

following graph.

Those sands designated as river sands on the graph often contain pebble sized lithic

fragments within the sand matrix. These units also exhibit a fining upwards sequence,



suggesting fluvial depositional conditions originally occurred in a high energy

environment capable of depositing large fragments (including possible floods depositing

angular lithics); that receded into a lower energy environment depositing fine grained

sands.

Figure 5.11 Moment skewness versus moment standard deviation to differentiate beach and river sands from the Stuarts Point aquifer (method adapted from Friedman, 1960). Numbers represent depth below ground surface for respective multi-levels.

5.2.2.2 The Onset Of Estuarine Conditions

There is wide debate in sedimentological literature over the accuracy of different

methods used to assess depositional environments based solely on grain characteristics.

As stressed in Leeder (1982) environments of deposition need to be deduced by a

combination of interpretative methods such as facies analysis, palaeocurrents and grain

size details. This is perhaps crucial to petroleum investigations, where sandstone

depositional environments are important in the exploration of petroleum resources. The



sediments of the Stuarts Point aquifer are unconsolidated and the study area is smaller

than most conventional petroleum investigations, however, a combination of

interpretative methods have been employed to satisfy any uncertainty present with the

use of a singular technique for deducing the depositional environment of the aquifer.

Thus far, analysis of the sand units has deduced that both barrier and fluvial

depositional conditions have contributed to the formation of the Stuarts Point aquifer.

The presence of clay units interbedded with sandy facies indicates an additional change.

Below the olive-grey fluvial sands (generally at an approximate depth of 18-20 m in

most boreholes) a gradual transition into a clayey sand is observed. The dual presence

of sand and clay suggests the high energy depositional environment of the fluvial and

beach sands was preceded by a lower energy depositional environment of mixed sand

and clay, and at deeper depths, solely clays. Such lower energy coastal environments

can include back-swamp lagoons, intertidal zones and estuaries. A technique was thus

sought to determine the environmental conditions that these clayey sand and

homogenous clay units had endured during deposition.

Numerous mollusc (shell) species were collected during drilling. Table 5.2 lists each

species identified herein, their location within the aquifer and their inferred habitat.

Depth ranges for mollusc occurrence within the aquifer have been provided in the table

below, however the cross sections in Figures 5.1-5.10 show the actual depths of each

identified fossil mollusc for each separate borehole.

The presence of various species at specific depths in the aquifer can suggest the likely

habitat endured by that mollusc and thus the depositional conditions occurring when the

surrounding sediments were deposited. Note that no molluscs were identified at depths

< 10 m in the aquifer; the same sands that have previously been identified as beach

barrier deposits. Therefore, the absence of mollusc species (particularly estuarine

species) in the shallow barrier sands supports the designation of shallow sediments

being deposited by onshore movement of sand in a high energy environment (i.e., beach

barrier formation).



Table 5.2 Mollusc species identified within the Stuarts Point aquifer. General depth of the mollusc fossil localities and their preferred habitats are also shown.

Species2 Habitat Stuarts Point

Locality (bore / depth)

Depositional environment and occurrence at Stuarts

PointPaphies elongata

(Reeve, 1854) Sandy beach with

exposed shores and high wave action

ML7/10.5 Beach10 m or less

Leiopyrga lineolaris

(Gould, 1861)

Sandy beach with exposed shores and

high wave action

ML7/10.5 Beach 10 m or less

Donax electilis (Iredale, 1930)

Littoral sand ML7/15 Sandy outer estuary zone

10-15 m Eumarcia fumigata(Sowerby, 1853)

Littoral (shoreline) sand

ML7/15ML7/25.5ML8/13.5

Sandy outer estuary zone

10-15 m 20-25 m

Bittium lacertinum (Gould, 1861)

Sandy estuarine beach with seagrass beds

ML7/16.5ML7/19.5

Sandy inner estuary zone

15-20 m Common ‘pipi’

FamilyHemidonacidae

Intertidal sandy beach ML7/23.5 ML8/25.5ML9/17.5

Sandy inner estuary zone

15-25 m Velacumantus

australis(Quoy and

Gaimard, 1834)

Seagrass beds ML7/19.5 ML7/25.5ML10/8ML10/11ML10/14ML10/21ML10/25ML10/30

Estuarine clayey sand > 10 m< 30 m

Anadara trapezia (Deshayes, 1840)

Subtidal estuarine mud ML7/19.5 ML7/25.5ML10/11ML10/14ML10/25ML10/30

Estuarine clay > 10 m< 30 m

Porthalotiacomtessi

(Iredale, 1931)

Sheltered bay with mud and sand

ML7/22 Estuarine clay > 20 m

Gari livida (Lamarck, 1818)

Sheltered bay with mud and sand

ML7/22 Estuarine clay > 20 m

Nodilittorina unifasciata

(Gray, 1826)

Intertidal (littoral) area on exposed rocks or mangrove trunks -common on rocky shores above high

water mark

ML8/26.5 Rocky shores on bedrock > 25 m

2 Brackets show the discoverer’s name and year of species discovery, rather than a citation to literature.



Sand designated as being deposited in the sandy outer estuary zone, and supported by

the presence of molluscs that live in such habitats, correspond to many sands previously

identified (Section 5.2.2.1) as being fluvially transported. It is suggested that these sands

were transported downstream by the Macleay River and deposited in the outer estuarine

moderate energy environment.

Below these sands, the transition to clayey units corresponds to species preferring sub-

tidal environments and seagrass beds. These clayey units are suggested to have

undergone lower energy depositional conditions in a tidal estuarine environment.

Mollusc identification of species collected in situ thus supports the designation of

estuarine conditions during deposition of the deeper parts of the Stuarts Point aquifer.

For comparison, studies conducted in nearby areas of the Macleay fluvial-deltaic plain

have also used identified mollusc species to define the depositional history of

sediments. Table 5.3 lists species that have been found in close proximity to Stuarts

Point.

Table 5.3 Common mollusc species and their environmental habitats identified in the sediments of the Stuarts Point aquifer.

StuartsPoint

Locality (bore / depth)

SpeciesSpecies Identification at other

Proximal LocalitiesReference for

Proximal Locality

ML7/15ML8/13.5

Eumarcia fumigata

(Sowerby, 1853)

12mbgs on the margin of the Macleay River fluvial-deltaic

plain

Hails (1964)

ML7/19.5ML7/25.5ML10/11ML10/14ML10/25ML10/30

Anadaratrapezia

(Deshayes, 1840)

9mbgs and 15-21mbgs in estuarine sediments at

Smithtown, 18km SW of Stuarts Point

Voisey (1934; Hails 1965)

ML7/19.5ML7/25.5ML10/8

ML10/11ML10/14ML10/21ML10/25ML10/30

Velacumantus australis

(Quoy and Gaimard, 1834)

0-10mbgs in the Macleay River fluvial-deltaic plain

Walker (1970)



For example, Walker (1970) identified a nearby estuarine unit that he termed the

Gladstone unit, which is suggested to have been deposited in the greater Stuarts Point

region as part of the Macleay fluvial-deltaic plain. Within this unit, Velacumantus

australis, a common mollusc living in seagrass beds, was used to designate the

sediments as being estuarine in origin. The same species was identified at Stuarts Point.

Clayey units within the Stuarts Point aquifer containing Velacumantus australis were

also designated as being deposited in an estuary. Further comparison to these local

studies will be made in Section 5.2.6.1 when all the evidence herein has been presented

and a new geomorphic model for the Stuarts Point aquifer has been proposed.

5.2.2.3 Weathered Bedrock

Several bores (ML1, ML6, ML9 and ML10) encountered hard rock refusal during

drilling and were terminated in suspected bedrock. ML9 encountered phyllite whereas

the remaining bores penetrated shale. The review of bedrock data presented in Chapter 2

(Section 2.3.2) shows inconsistent theories relating to the bedrock lithology at Stuarts

Point. From the compilation and review of this data, in addition to borelog analysis

presented here, bedrock for the Stuarts Point area has been deduced for this study and

shown in Figure 5.12. As with most other geological mapping investigations carried out

in this area, these boundaries are simply inferred until more accurate data is obtained.

The northern part of the aquifer is underlain by Lower Palaeozoic Nambucca (Pee Dee)

Beds consisting of slate and phyllite and intruded by the Yarrahappinni coastal

granitoid. These beds were intersected at the base of ML9, which encountered phyllite.

The overlying Kempsey Beds extend from immediately south of the Stuarts Point

township to approximately 9436-2-N EUNGAI 001 885 and consist of mudstone.

Bedrock was not encountered within this area and so lithology is proposed based on the

literature review.

South of this approximate grid reference, bores penetrated shale bedrock. There are two

main possibilities for this change in lithology. The Kempsey Area transcurrent fault

diagonally separates the mudstone of the Lower Palaeozoic Kempsey Beds from the

Upper Palaeozoic (Permian) shale of the Macleay Series; or, contact metamorphism

associated with the granitoid intrusion does not extend this far south, leaving the shale



beds unchanged for the southern bores but metamorphosing shale into phyllite in close

proximity to the intrusion.

Figure 5.12 Proposed bedrock units at Stuarts Point. The Stuarts Point aquifer is contained within the dotted area. Approximate bedrock contours (based on drilling evidence) are shown at 5m depth intervals.

The approximate depth to bedrock at Stuarts Point is uncertain. In 1982, the Water

Resources Commission (WRC) was engaged to assess the feasibility of a proposed well

field southwest of the Stuarts Point township. The use of previous mining borelogs

complemented with logs from three WRC test bores provided a rough bedrock contour

map. No details were provided on stratigraphic units encountered (i.e., Nambucca Beds

and/or Kempsey Beds), however the lithologies that were penetrated were recorded. By

combining the data obtained by WRC in 1982 with this most recent data, probable

bedrock contours have been reassessed and placed on the proposed bedrock diagram

shown in Figure 5.12.



Depth to bedrock can give an indication of aquifer thickness. Bedrock refusal was first

encountered at 22 mbgs (ML9) close to Mt Yarrahappinni. It is assumed the aquifer

shallows as it progresses northward toward the mountain. The deepest bore used in this

study, ML6, encountered bedrock at 40 mbgs and is located in the southern section of

the aquifer. Anecdotal evidence suggests bores drilled in the centre of the aquifer extend

past 40 m depth before encountering bedrock.

Combined with the uncertainty of bedrock structure and lithology, the actual thickness

of the Stuarts Point aquifer is suggested to range from shallow (a few metres thick near

Mt Yarrahappinni) to deep (perhaps up to 60 m depth in the centre of the aquifer) and

moderate (30-40 m deep in the southern sections of the aquifer). Where drilling

encountered weathered bedrock and clay, no mollusc fossils or mixed sand facies were

encountered, thus differing to the estuarine-deposited clays. More than often, the

bedrock clays were the same colour as bedrock and contained fragments of weathered

rock, distinguishing them from estuarine clay deposits.

5.2.3 Aquifer (Sediment) Chemistry

A full examination of aquifer composition was conducted during the course of these

investigations. Thirty-six sediment samples from multi-levels ML7-ML10 were

analysed for total element content every 3-4 m depth by XRF. Table 5.4 shows mean

concentrations of major and trace elements in the Stuarts Point sediments. Full results

are provided in Appendix B1.

5.2.4 Aquifer Facies Definition

5.2.4.1 Statistical Procedures

Statistical analyses were conducted on sediment chemistry to ascertain if solid phase

chemistry could indicate source provenance or aquifer depositional variability. Thirty-

six sediment samples were subjected to Hierarchial Cluster Analysis (HCA), clustered

according to total element XRF compositional similarities observed between samples.

In total, 33 elements were involved in the statistical analysis. Appendix B1 lists the full

chemical data set. Two dominant clusters (Clusters 1 and 2) with sub-groups (1A and



1B, 2A and 2B) were differentiated on the resulting dendrogram shown in Figure 5.13.

These clusters link the sediments that are most similar based on chemistry.

Table 5.4 Element means in the Stuarts Point aquifer sediments, as determined herein by XRF (n=36).

ElementMean Reported in Stuarts

Point Sediments Majors expressed as % oxides

SiO2 88.2 Al2O3 4.10 TiO2 0.30

MnO2 0.01 Fe2O3 1.30 MgO 0.40 CaO 1.50 K2O 0.80

Na2O 0.60 P2O5 0.05

Trace elements shown in mg kg-1

As 6.30 Ba 178 Ce 35.2 Cd 0.20 Co 32.8 Cr 22.3 Cu 13.3 Ga 5.80 Mo 1.00 Nb 3.40 Ni 7.70 Pb 7.30 Rb 36.1 Sb 6.00 Sn 0.00 Sr 76.7 Th 3.20 U 1.80 V 48.6 Y 8.50 Zn 21.3 Zr 94.5

Sediment samples belonging to Cluster 1 are dominated by sand and occur in the top

15m of the aquifer. Samples belonging to sub-cluster 1A exhibit a fining upwards

sequence of sand from approximately 6-15 mbgs. These sands represent fluvially



transported sediments and correspond with the statistically defined river sands from

Figure 5.11. Lenses of these “fluvial sands” also occur at various other depths in the

aquifer representing the ever-changing course of the Macleay River. Sands in sub-

cluster 1B are fine to medium grained homogenous sands existing from the surface to

approximately 6 mbgs. These sands have been deposited via onshore sediment supply

and correspond to the statistically defined “beach barrier deposits”.

Sediments belonging to Cluster 2 are a mix of sand, silt and clay. Sub-cluster 2A

sediments contain the highest amount of clay material and represent clays encountered

towards the base of the bore. These clays are therefore deemed to be a product of

bedrock weathering and thus are statistically grouped together based on their

mineralogical and textural composition and the weathering effects of the underlying

bedrock. Sub-cluster 2B represents sands that are influenced by both fluvial processes

but also contain some clay, thus being termed “fluvial sand/estuarine clay” sequences.

These sediments exist from 15 mbgs to the “bedrock clays” at the base of the bore.

The results of the statistical clustering show that the “fluvial sands” and “beach barrier

deposits” are closely related (Euclidean distance of 7); while the “bedrock clays” and

“fluvial sand/estuarine clay” facies are also similarly related to one another (Euclidean

distance of 7). Note that two populations exist within sub-group 2B, which effectively

separates the fluvial sands and estuarine clays of ML7 from the same proposed facies in

other bores. This is likely a product of the close proximity of ML7 to the current

estuary, thereby exposing ML7 to more recent geomorphic activity. Given that

sediment chemical composition can be largely based on grain size parameters (i.e., clays

and silts provide more surface area for element sorption than sands) it is postulated that

assignment to each cluster group may be highly influenced by chemical differences

induced by grain size characteristics, rather than being representative of their source

environment. However, the statistical analyses have successfully aided in the

differentiation of facies within the aquifer, particularly when supported by the results

presented so far, as discussed in the next section.



Figure 5.13 The HCA dendrogram on sediment chemistry successfully delineated the proposed facies units.



5.2.4.2 Development of Aquifer Cross Sections

The utilisation of grain size parameters, identification of fossil molluscs, depth to

bedrock and statistical grouping of sediment chemistry, enabled the construction of

cross sections in Figure 5.14 to Figure 5.17.

Sands differentiated by grain size analyses of skewness and sorting have been

stratigraphically termed “fluvial sands” or “beach barrier deposits” and have been

plotted on the cross sections. Individual samples assigned to either beach (B) or river

(R) sections of the graph in Figure 5.11 are also annotated on the cross sections.

Facies containing mollusc fossils helped to identify specific depositional environments.

The results of the statistical clustering have also been placed on the cross sections as the

final piece of evidence supporting the designation of separate facies within the aquifer.

In summary, these cross sections successfully show the following sediment facies exist

within the Stuarts Point aquifer:

“Beach barrier deposits” – sands deposited by onshore movement during

stillstands

“Fluvial sands” – deposited by the Macleay River

“Fluvial sand / estuarine clays” – deposited by river and tidal movement

“Bedrock clays” –representing weathered bedrock

Figu

re 5

.14

Cro

ss S

ectio

n A

to A

'. Se

e te

xt fo

r di

scus

sion

.

Figu

re 5

.15

Cro

ss se

ctio

n B

to B

'. Se

e te

xt fo

r di

scus

sion

.

Figu

re 5

.16

Cro

ss S

ectio

n C

to C

'. Se

e te

xt fo

r di

scus

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.

Figu

re 5

.17

Cro

ss se

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n D

to D

'. S

ee te

xt fo

r di

scus

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.



5.2.5 Determining Source Provenance for the Aquifer Facies

5.2.5.1 Origin of the Sands

The previous sections have successfully differentiated between different sediment facies

within the aquifer, suggesting that deposition of the aquifer was completed under

different geomorphic conditions. Before a full geomorphic history is proposed for the

Stuarts Point aquifer, some information on the possible source provenance of the aquifer

sediments is required.

The evidence presented so far indicates that a significant portion of aquifer sediments

were deposited by fluvial processes. Consequently, the original provenance of these

fluvial sands is considered highly plausible as being derived from the regional

catchment of the Macleay River and its geologic base; the NEFB. Large lithic fragments

observed in many of the fluvial sediment samples are characteristically composed of

metamorphics, assumed to be derived from the NEFB. However, a second approach was

utilised to confirm, or deny, the hypothesis that the aquifer sands were originally

sourced from the NEFB3.

Ternary diagrams are commonly used by sedimentologists to determine sedimentary

rock classifications. The relative abundance of quartz, feldspar and lithic fragments

present in the sample enables the rock to be named accordingly. Additionally,

determining the source provenance for sediments is common in petroleum

investigations. A ternary diagram can be divided into the tectonic settings seen in

Figure 5.18. The location of the sediment samples on the diagram can indicate if the

sediment is derived from an arc, basement uplift, or a recycled orogen provenance.

Analysis of the major mineral composition (quartz, feldspar and lithics) of the Stuarts

Point aquifer sediments was estimated and plotted on a Ternary diagram (Figure 5.18)

to indicate the source provenance for the aquifer sands. Actual estimates of each

respective mineral abundance is tabulated in Appendix B4.

3 N.B. Ternary diagrams are commonly used in sandstone petroleum investigations and thus provide an indication only of the aquifer sand provenance.



Figure 5.18 Ternary diagram of the Stuarts Point sediments showing a recycled orogen for most aquifer sediments. Boundaries for tectonic fields are from Dickinson and Suczek (1979), Dickinson et al. (1983), and Ingersoll and Suczek (1979) (Q=quartz, F=feldspar, L=lithics).

Both “fluvial sands” and “beach barrier deposits” plot within the recycled orogen

portion. The diagram reflects their high quartz content, particularly the beach barrier

deposits, which are very well sorted and homogenous. These sands have undergone

significant transport and re-working. They were most likely derived from the NEFB and

transported off-shore, before deposition of the fluvial sands occurred. The fluvial sands,

although quartz rich, also contain some lithic fragments, assumed to be derived from the

NEFB. They would likely evolve towards quartz-rich sands should they undergo a

similar re-working and transport process as the “beach barrier deposits”. Both sand units

contain minor amounts of feldspar derived from the weathering of the NEFB.



5.2.5.2 Selection of a Hinterland Indicator Element

To ascertain a better indicator of source provenance the regional mineralisation has been

examined. The aim of this section is to select an element that is abundant in the upper

reaches of the Macleay river, in addition to being capable of fluvial transport by the

river, deposition within the aquifer sediments, and being retained within the aquifer

matrix. Such an element can then act as an ‘indicator’ of a NEFB provenance for the

aquifer sediments. Trace elements associated with rare ore deposits, may act as potential

NEFB hinterland indicators.

Given its geological complexity, the NEFB contains a wide variety of mineralisation,

each type capable of contributing trace elements to the aquifer composition:

The accretionary prism sediments associated with subduction related complexes

in the mid-Palaeozoic era, carried with them syngenetic mineralisation and the

source for later epigenetic mineralisation, particularly gold and antimony veins

(Gilligan and Barnes, 1990).

After a period of intense folding and faulting, mid-Triassic granites intruded

parts of the NEFB to produce gold, molybdenum, bismuth and tungsten deposits

(Suppel, 1974) and minor elements associated with these granitic intrusions and

their contact aureoles.

Tertiary volcanism produced sapphires.

Placer deposits of tin, gold, sapphires and diamonds formed in the Tertiary and

Quaternary, with the Quaternary also producing heavy mineral sand deposits

along the eastern coastline.

A large river draining this mineralised hinterland can deposit sediment in its lower

reaches characterised by a geological signature of its prior incision pathway. The

Macleay River is the second largest river on the NSW north coast. Figure 5.19 shows

the various mineral deposits, hence potential hinterland indicators, located in the upper

Macleay River catchment.



Figure 5.19 Indicative types of mineralisation occurring within the Macleay River catchment (information obtained from Gilligan et al., 1992).

Numerous arsenic deposits are located within the upper reaches of the Macleay River

catchment, however the most abundant form of mineralisation belongs to the antimony

containing mineral, stibnite (Sb2S3).

Data on antimony cycling in the environment is relatively scarce. Filella et al. (2002a;

2002b) provide comprehensive reviews of current knowledge on antimony processes to

date. They do, however, suggest more research is needed to confirm the current findings

and deduce the full biogeochemical cycle of this element. There is a lack of information



on antimony abundance in sediments. In their review, Filella et al. (2002b) agreed that

existing data seems to indicate antimony in soils and sediments is associated with

relatively immobile phases, thus making it a good hinterland indicator as it should have

survived the fluvial transportation process.

Elevated concentrations of antimony in sediments is either anthropogenically related or

associated with high arsenic concentrations in sulfidic ores (Filella et al., 2002a).

Mitchell and Burridge (1979) report a crustal mean of 0.2 mg kg-1. Soils in the United

States have a reported geometric mean of 0.48 mg kg-1 and range of <1 – 10 mg kg-1

(Adriano, 1986). Contaminated sediments from a historical mining area in Germany

ranged from 13 - 1,317 mg kg-1 Sb (Hammel et al., 1998).

Mean Sb concentrations in the Stuarts Point sediments are 6 mg kg-1 (Table 5.4), more

than 12 times the mean for soils in the United States. In addition, their persistence in an

aquifer dominated by sand is unusual, since sand normally does not contain numerous

sites for adsorption (unless dominated by oxide coatings). Clays, however, do. An

analysis of Sb distribution within the proposed aquifer facies is provided in Table 5.5.

Table 5.5 Mean Sb concentration for each of the proposed aquifer facies at Stuarts Point.

Aquifer Facies Mean Sb (mg kg-1)Beach Barrier Deposits 9.80

Fluvial Sands 4.73 Fluvial Sand / Estuarine Clays 4.71

Bedrock Clays 3.55

The “beach barrier deposits” contain the highest reported mean Sb concentrations,

possibly in an immobile oxide form. Since both the fluvial sands and the barrier sands

are suggested by Figure 5.18 to be derived from a recycled orogen provenance, perhaps

the beach sands were eroded and transported at a time when natural erosion of stibnite

deposits in the hinterland were accelerated. This may occur as the river changes course,

exposing and eroding a mineral deposit directly, and then moving away from the deposit

or eroding its entire supply.



The “fluvial sands” and “fluvial sand / estuarine clays” contain similar Sb

concentrations, showing the fluvial depositional conditions were intermittent with

estuarine conditions. The “bedrock clays” at Stuarts Point exhibit low concentrations of

Sb due to their genesis from in-situ bedrock which is unlikely to contain stibnite

deposits. In summary, the high concentrations of Sb in the aquifer sediments,

particularly fluvial sands and beach sands derived from the same provenance, suggests

that the majority of sediments in the Stuarts Point aquifer have been derived from

erosion of the NEFB mineralised hinterland.

5.2.6 Aquifer Geomorphology

Initially, the author supported the classification of the Stuarts Point sediments as a dual

barrier system, proposing the model of Roy and Thom (1981) could be applicable to

Stuarts Point (O’Shea and Jankowski, 2002). Strong evidence now suggests that the

majority of aquifer sediments have been sourced from the hinterland. The designation of

appropriate stratigraphic units in the Stuarts Point aquifer has been successful and in

addition, these units correspond with research conducted by other scientists in nearby

localities.

5.2.6.1 Linking the Stuarts Point facies to surrounding depositional environments

The Macleay River drains hilly tablelands upstream of Kempsey at which point it

broadens into a wide fluvial-deltaic flood plain termed the Lower Macleay valley. The

Stuarts Point aquifer forms part of this floodplain.

Walker (1970) examined exposed soil terraces upstream of Kempsey and linked several

terraces to buried floodplain deposits downstream, confirming the input of fluvial

sedimentation within the Lower Macleay floodplain / Stuarts Point aquifer. An analysis

of Walker’s (1970) floodplain stratigraphy provided the following comparisons to the

Stuarts Point sediments (from oldest to youngest):

The older clays found at the base of most bores and termed “bedrock clays” herein,

are comparable to Walker’s stiff sands and clays and basal clay occurring towards

the bottom of the floodplain profile. Walker (1970) could not successfully link this

unit to upstream soil terraces, and additionally, none of the bores examined in his



study penetrated bedrock. Intersection of bedrock during the investigations herein

indicates this basal clay is likely to be a product of in-situ weathered bedrock rather

than transported sediment.

Walker’s main unit comprising the Macleay River floodplain is termed the

Gladstone deposit; a thick (up to 23 m) estuarine unit containing the same shell

species (Velacumantus australis, Quoy & Gaimard) as those identified within the

“fluvial sands / estuarine clays” denoted herein (Section 5.2.2.2). The Gladstone

deposits were described by Walker as being oxidised in the top layers and

indicative of currently termed acid sulfate soils. A wood sample collected from the

base of the Gladstone deposit has been dated as 8,530 +/- 200 years. Sandy deposits

are more common towards the east in the vicinity of Stuarts Point, providing

support to the combination of “fluvial sands / estuarine clays” found within the

study area.

Located stratigraphically below and above Walker’s Gladstone deposits are fluvial

sands with coarse point-bar deposits. Those sands overlying the Gladstone deposits

contained a wood sample dated as 3,295 +/- 95 years, indicating a change in

depositional environment from estuarine to fluvial. This date potentially

corresponds to a time when sea level was at a stillstand along the eastern coast of

NSW. These sands are termed “fluvial sands” herein and represent the former path

of the Macleay River.

Since no age dating was conducted as part of these investigations, the ability to correlate

the Stuarts Point facies with nearby depositional environments, combined with the

literature available on sea level movements in the Quaternary, allows the following

geomorphic model to be proposed for the deposition of the Stuarts Point aquifer.

5.2.6.2 Proposed Geomorphic Model for the Stuarts Point Aquifer

During Early Pleistocene times (>18,000 B.P.) sea level was approximately 24-30 m

higher than its present level (Hails, 1968) and the coastline was located inland from the



current coastline at the base of rocky outcrops belonging to the NEFB. Coastal granitoid

intrusions were present as offshore islands (Figure 5.20a).

A drop in sea level (–135 m) consistent with that observed in Bangladesh (Acharyya et

al., 2000) during the Late Pleistocene (~18,000 B.P.) exposed the underlying bedrock

connecting the granitoid intrusions to the mainland (Figure 5.20b). These low sea levels

were conducive to valley incision and fluvial sedimentation. Coarse sediments including

gravels were deposited by the river.

A slight rise (+4-5 m) in sea level soon after (18,000-12,000 B.P.) caused the coast to

advance landward (Figure 5.20c). A constant supply of sediment onshore led to the

formation of the inner Pleistocene sand barrier as suggested by Hails (1968). Still in the

Late Pleistocene (12,000-10,000 B.P.), a temporary regression saw the coastline retreat

leading to the oxidation of exposed sediments (Figure 5.20d) similar to weathering and

oxidation observed in parts of the Ganges delta (Acharyya et al., 2000). Fluvial gravels

in oxidised silty-clay, indicative of this regression, were identified at approximately 26

mbgs in bore ML10.

In the Early Holocene (10,000-7,000 B.P.) sea level rose once more (Acharyya et al.,

2000; Chappell and Polach, 1991) leading to fluvial and estuarine conditions

overlapping as both river and sea continually changed their course. The Macleay

fluvial-deltaic plain expanded as the river deposited its sediment load on its journey to

the sea.

In more recent times (since ~7,000 B.P.) the sea level has stabilised (Williams et al.,

1998) and formation of the outer Holocene barrier (Hails, 1968) continues to occur via

onshore sediment supply (Figure 5.20e). Thus the evidence presented herein suggests

that a combination of fluvial, estuarine and onshore sediment supply have each

contributed to aquifer formation under conditions similar to those proposed for the

deposition of alluvial aquifers with high-As groundwaters in Bangladesh.



Figure 5.20 Proposed geomorphic model for the deposition of the Stuarts Point aquifer.



5.3 DISCUSSION ON EACH PROPOSED ARSENIC SOURCE HYPOTHESIS

5.3.1 Anthropogenic Sources

Its relatively low population and small town status do not exclude Stuarts Point from the

possibility of anthropogenic contamination sources. Its fertile land and quality natural

resources are among the main reasons Stuarts Point is a habitable area. Agriculture is

the dominant land use in the area, but numerous other potential arsenic contamination

sources are present within or near the Stuarts Point aquifer (Figure 5.21).

Figure 5.21 Potential anthropogenic arsenic sources in the Stuarts Point aquifer.



5.3.1.1 Agricultural Practices

Aldarman et al. (1984) provided a folk history of Yarrahapinni and Stuarts Point, which

is briefly drawn upon here to examine potential past arsenic contamination practices.

Early settlement industries at Stuarts Point included timber milling, sea transport and

fishing before dairy farming became common in the late 1800’s. By 1984 beef cattle

had replaced dairy farming. In 1920 the first banana plantations were established and

continue to be a dominant land use today. Other crops previously grown have included

asparagus, watermelon, corn, pumpkin and beans. Current crops include potatoes, stone

fruits, avocados and macadamias.

Both cattle and crop farming practices have the potential to provide sources of arsenic to

the environment. A common form of cattle tick eradication involved the use of arsenic

dipping solutions which were common on the NSW mid-north coast from 1895 until

replacement with DDT in 1955 (Maganov et al., 2000). Studies by Kimber et al. (2002)

of former cattle dip sites in NSW showed no signficant transport or redeposition of

arsenic had occurred at the sites studied (n=28). Sandy soils exhibited groundwater with

elevated arsenic concentrations, however, these declined sharply with increased distance

(ca. 20 m) from the dip sites studied. NSW Agriculture retains records of former cattle

dip sites located throughout the State and does not posses any documentation of such

sites being present at Stuarts Point.

Arsenical herbicides and pesticides, such as sodium arsenite, have been widely used in

the past and continue to be used around the world today. No documentation on the types

of pesticides/herbicides previously used at Stuarts Point is available, however

application is assumed to have occurred, particularly on banana plantations. Arsenic

may or may not have been an active ingredient in the pesticides/ herbicides used.

Fertilizers applied to assist crop growth can contain up to 6,200 mg kg-1 of arsenic

depending on fertilizer type (eg. micronutrient, phosphate) and manufacturer

(Woltering, 2004). An inventory of all fertilizer types used at Stuarts Point is not

available but the use of Muriate of Potash (potassium chloride) has been confirmed. In a

study investigating trace element concentrations in this type of fertilizer Robarge et al.



(2004) reported arsenic concentrations less than detection limit (0.3 mg kg-1) for all

samples analysed (n=8).

The application of fertilizers, pesticides and herbicides has the potential to be diffuse

sources of arsenic contamination within the aquifer. Some degree of traceability would

be expected in the hydrochemical data and may aid in the identification of any diffuse

arsenic sources in the aquifer. No such patterns have been identified during these

investigations.

5.3.1.2 Historical Mining

The east coast of Australia contains valuable heavy mineral sand resources (rutile,

zircon, monazite) and Stuarts Point is no exception. In the 1970’s localised mining

occurred to the southwest of the township upgradient of all multi-levels sampled for this

study. Separation processes between heavy mineral fractions and other grains is

normally based on physical separation methods, dismissing the use of any chemicals

during the mining process and reducing the production of contaminated by-products.

Sand mining occurring within the Tomago Sandbeds on the mid-NSW coast south of

Stuarts Point (Newcastle) has been tentatively linked to arsenic release into the

groundwater environment via the oxidation of arsenic-rich pyrite. In contrast to Stuarts

Point, groundwater in the Tomago Sandbeds contains elevated levels of iron indicative

of pyrite oxidation which decreases away from the mined area (Binnings et al., 2001).

No such impact is evident downgradient of the mined area at Stuarts Point.

A small unoperational (?) quarry located in a hillside cutting west of the township is

presumed to supply slate and phyllite for use in tiling and decorative construction and is

not considered to impact the quality of the groundwater.

5.3.1.3 Town (Anthropogenic) By-Products

Kinki Cemetery has been irregularly used since 1877 and has significant European

Heritage value (NSW DPWS, 1998). In a study examining inorganic soil contamination

from cemetery leachate, Spongberg and Becks (2000) found statistical differences

existed between arsenic concentrations in soils at a (different) cemetery site (1 to 20 m

from the nearest grave) compared to soils tested off the cemetery site. The elevated



arsenic concentrations – up to 8.06 mg kg-1 – were suggested to be a product of previous

embalming practices and/or coffin wood preservatives. Nearby graves dated back to the

mid-late 1800’s in an era when arsenic was commonly being used in the preservation

process. Operation of Kinki Cemetery since 1877 lends support to the possibility of

embalming fluids being a potential source of arsenic to the Stuarts Point environment.

The use of arsenic in embalming fluids was apparently stopped in 1910 (Dent, 2002).

Prior to this, the use of such embalming fluids was minor in Australia (Dent, pers.

comm., 2004).

The limited studies conducted on contaminants leaching from cemeteries (Dent, 2002;

Spongberg and Becks, 2000) indicate arsenic leaching from embalming fluids and/or

treated coffins may be localised to within the cemetery proximity, however further

information is required to confidently support these conclusions, and notably, soil and

groundwater conditions vary between cemetery sites. The potential for Kinki Cemetery

to contribute widespread arsenic contamination within the Stuarts Point aquifer is

therefore considered unlikely.

Prior to the installation of a reticulated water supply at Stuarts Point in 1985, concern

was raised over the impact of septic discharge contaminating spearpoints utilised as

domestic water supplies (NSW DPWS, 1998). In 1999 DLWC informed local residents

that uncertainty still remained over the impact of septic tanks on the local groundwater

system. A study was initiated to examine the effects of both the local rubbish tip and

septic tank systems on the groundwater environment, but to date has not been released.

Investigations conducted herein do not show any relationship between these potential

contaminant sources and arsenic distribution throughout the aquifer, thus dismissing

their probability as arsenic sources within the aquifer.

The treatment plant recently installed to remove arsenic from the town water supply is

obviously not the original source of arsenic and stringent controls are in place to contain

the arsenic rich by-products produced during the treatment process, so as to ensure

further contamination does not occur.



The NSW Environmental Protection Authority (EPA) maintains a Contaminated Sites

Register throughout NSW. The database includes all sites that have been deemed

contaminated regardless of remedial measures that may have since been implemented.

There were no listings for Stuarts Point, Mt Yarrahapinni, Clybucca or Fishermans

Reach on the Contaminated Sites Register as of late 2004.

5.3.2 A Regional geologic source?

In the absence of any notable anthropogenic arsenic source, the focus turns to natural, or

‘geogenic’ sources. Past investigations (Section 2.4) have largely based their source

hypotheses on the previously accepted geomorphic model for the aquifer. Since the

commonly accepted geomorphic model for the aquifer was based on barrier

development from onshore sediment transport (Section 2.3), the possibility of arsenic

being derived from the hinterland has not been fully examined.

Section 5.1 now proposes that a significant amount of the aquifer has, in fact, been

derived from erosion of the hinterland. The presence of elevated Sb in the aquifer

sediments, supplied from erosion of abundant stibnite deposits in the NEFB hinterland,

supports this hypothesis. Of more notable consideration is that natural stibnite deposits

can contain more than 5,000 mg kg-1 arsenic in solid solution, in addition to its common

association with arsenopyrite and arsenian pyrite (Ashley et al., 2003).

A geochemical study into the environmental mobility of Sb and As from mesothermal

stibnite deposits in a high energy stream environment undergoing accelerated natural

erosion, was conducted by Ashley et al. (2003) in the upper reaches of the Macleay

River catchment (Hillgrove – see Figure 5.19). They found that significant levels of Sb

and As had been contributed to the Macleay River system through both mining

activities and natural processes. It is noted that the Stuarts Point sediments contain up

to 14 mg kg-1 of Sb (Section 5.2.5.2; Appendix B1) and are thus considered Sb-rich.

These concentrations are comparable to As concentrations, which are also up to 14 mg

kg-1 (Chapter 6; Appendix B1) thus both elements are considered to be elevated in these

sediments. It is possible the two elements may exhibit similar properties given their

proximal locations in the periodic table, however, the relative immobile nature of Sb is

expected to override any non-geochemical relationship.



The approximate 1:1 Sb/As concentration ratio also corresponds to the 1:1 Sb/As ratio

observed by Craw et al. (2004) in mine rocks and waters from the Hillgrove antimony

mine located in the upper reaches of the Macleay River catchment. A statistical

correlation4 of Sb and As concentrations in the Stuarts Point aquifer sediments produces

a moderate negative correlation (r2 = -0.58) between As and Sb (Figure 5.22). For

comparison, the only other negative correlation with arsenic was SiO2 (r2 = -0.54) and

the highest positive correlation was with Fe oxide (r2 = 0.67) (Appendix B7 shows full

statistical correlation results; Appendix B9 shows XY plots).

Figure 5.22 Sb versus As concentrations in the aquifer matrix. The “beach barrier deposits” contain the most Sb and the least As, showing Sb is more immobile than As.

4 Pearsons correlations performed in SPSS (Appendix B6).



The predicted 1:1 ratio (observed upstream at the minesite) has potentially been altered

during erosion, transport and the subsequent onset of geochemical conditions in the

aquifer. A positive correlation of 1 could be expected if both elements behaved

conservatively during transport. Hammel et al. (1998) found that antimony in sediments

derived from geogenic sources is mostly immobile. The moderate negative correlation

indicates Sb is less mobile than As which may contribute to the onset of dissolved As

concentrations in the groundwater.

An independent and concurrent study examining Sb and As enrichment in the surface

deposits of the Macleay floodplain assumes a predominant source of both Sb and As

from upstream dispersion (Tighe et al., 2005). While their calculated Sb and As

enrichment in the top 10 cm of floodplain soil agrees well with estimates lost from

upstream mining areas (Ashley and Graham, 2000), results reported herein show this

enrichment extends to depths of more than 20 m when sediments were deposited prior

to mining activities. Discussions with these researchers (Tighe, pers. comm., 2005)

reveal that preliminary work conducted by the University of New England indicates

shallow soil Sb enrichment has most likely occurred after commencement of mining at

Hillgrove; but that Sb concentrations at depth in these floodplain deposits may be a

product of natural erosion prior to mining. These combined results indicate the natural

erosion of Sb and As has potentially been occurring for several thousand years.

It is therefore proposed that the source of arsenic within the aquifer may be of regional

geological origin, primarily stemming from the erosion of hinterland stibnite deposits.

The original mode of transport for arsenic may be as a sorbed constituent, or

alternatively, present in dissolved form. After deposition of the aquifer sediments,

antimony is retained by the solid phase whereas arsenic is prone to mobilisation. This

causes the negative correlation between the two elements. It is suggested that antimony

is retained in the sediments in an immobile oxide form, however further work is

required to confirm or deny this hypothesis.

5.3.3 Input from As-rich Holocene Sea Level Rise

Smith et al. (2006) suggested the Stuarts Point groundwater was elevated in arsenic due

to adsorption onto clays during sea level fluctuations in the late Pleistocene and



Holocene (Section 2.4). They support their theory by reference to several studies

(Acharyya et al., 2000; Khan et al., 2000; Manning and Goldberg, 1997; and Lin and

Puls, 2000).

The first two studies indicate a correlation between arsenic and Bangladeshi sediments

subjected to Holocene sea level fluctuations. It appears that Smith et al. (2006) are

suggesting that since the Stuarts Point aquifer has also endured sea level movements in

the Holocene, then the transgressing sea must be the source of arsenic in both aquifers.

Indeed, the Stuarts Point aquifer has potentially been subjected to the same eustatic

movements in sea level as the Bengal Delta Plain5.

Table 5.6 compares the two alluvial aquifers with respect to sea level change and

sediment depositional environment.

One critical difference is observed between marine transgressions occurring in

Bangladesh when compared to Stuarts Point. Khan et al. (2001a) stated;

“The major source of arsenic in the Quaternary sediments of the Bengal Delta is

inferred to have derived from arsenic-rich aqueous solution due to marine

transgression.” (pg 57)

In other publications (Khan et al., 2000; 2001b) the arsenic-rich aqueous solution is

described as seawater transgressing over volcanic ash deposits rich in arsenic, thereby

concentrating it within the transgressing seawater. As sea levels later regressed, the

arsenic was retained by iron and manganese oxyhydroxides present within the fine-

grained aquifer sediments. The critical difference is that no such arsenic-rich deposits

have been identified at Stuarts Point.

5 Hails (1968) states that while eustatic events during the Quaternary were likely to be worldwide, comparison between different hemispheres must be made with caution due to the common lack of evidence of former shorelines in some coastal regions.



Table 5.6 Summary of eustatic changes in sea level and comparative depositional environments between the Bengal Delta Plain and the Stuarts Point aquifer.

Epoch a

(YearsAgo)

Sea Level Movement

Depositional Conditions in the Bengal Delta Plain b

Depositional Conditions at Stuarts Point c

EarlyPleistocene (>18,000)

High sea level ~ +24-30m

Sea level inundation Pre-Pleistocene shoreline of Hails (1968) with granitoids as

offshore islands Late

Pleistocene(18,000)

Sea level at its lowest level ~

-135m

Shoreline at location of outer continental shelf;

Pleistocene & Late Tertiary sediments exposed; valley

incision & fluvial (basal sand/gravel) sedimentation

Offshore islands connected to the mainland; Permian

metasediments and sedimentary bedrock layers exposed; valley

incision & fluvial (basal sand/gravel) sedimentation

LatePleistocene

(18,000-12,000)

Sea level rise ~

+4-5m

Silts and clays were deposited over the eroded

Pleistocene or Upper Tertiary sediments and filled

entrenched fluvio-deltaic valleys

Formation of the Pleistocene inner barrier

LatePleistocene

(12,000-10,000)

Temporaryregression

Weathering/oxidation of parts of the Ganges delta

Exposure and oxidation of fluvial gravels

EarlyHolocene(10,000-7,000)

Sea level rose again

Flooding (low energy) of the entrenched valley courses of the rivers; conversion to fluvial marshes, lagoons and estuaries; swampy

environments conducive to mud/silt deposition

Fluvial & Estuarine conditions overlap as sea level and the

Macleay River course continually change; deltaic

plains constructed landward of barriers

Holocene(7,000-5,000)

Rapid rise in sea level

Further invasion by tidal mangroves; extensive

development of marine & freshwater peats

Continuation of above

Holocene(5,000-current)

Slow or continuous rise

in sea level followed by a

stillstand

Migration of the shoreline to present day level

Formation of the Holocene outer barrier; current shoreline

conditions

a Dates are indicative only b Acharyya et al. (2000)c Hails (1968) and herein



Smith et al.’s (2006) second supporting reference, Acharyya et al. (2000), also noted a

positive correlation between arsenic-bearing zones in Bangladeshi aquifers with the

fluvial-deltaic alluvium subjected to Holocene marine transgression. However, in

contrast to proposing a seawater arsenic source, they suggested arsenic was likely to be

derived from the decomposition of arsenic-bearing sediments in the source area (i.e.,

hinterland) which was then subsequently adsorbed by secondary oxyhydroxides.

Ravenscroft (2001) agreed that the origin of arsenic in Bangladesh was geological,

indicating there were numerous sources of arsenic in the hinterland, but that the critical

components were the weathered minerals contributing oxyhydroxides to the suspended

load of rivers where arsenic could later be adsorbed. This is supported by Appelo et al.

(2002) who explain that high arsenic (2-29 g L-1) is transported via river water in

Bangladesh allowing much arsenic to sorb to iron oxides in the river sediments resulting

in arsenic available in aquifers built from these sediments. These latter theories are

consistent with a hinterland arsenic source as proposed herein for Stuarts Point.

Acharyya et al. (2000) also note that preferential adsorption of aqueous arsenic into

fine-grained sediments occurred under low energy regimes (i.e., those sediments

deposited during sea level transgression in the Holocene in Bangladesh). Figure 5.23

shows arsenic solid phase concentrations peak within the Stuarts Point “fluvial

sand/estuarine clays” before decreasing towards the base of the bore and the “bedrock

clays”. As is the case in the Bengal Delta Plain, the finer grained sediments provide

more surface area for arsenic adsorption and thus it is expected that the clay units within

the Stuarts Point aquifer would retain the most arsenic, regardless of arsenic source.

Given the lack of an aqueous arsenic source (i.e., dissolved volcanic ash deposits) to

enrich the transgressing sea, the Stuarts Point aquifer is unlikely to have been inundated

by an arsenic-rich aqueous solution. The average concentration of arsenic in seawater is

1.5 g L-1 (Smedley and Kinniburgh, 2002) and it is assumed this was the amount

present in the transgressing seawater endured by the sediments at Stuarts Point.



Figure 5.23 Vertical solid phase arsenic concentrations reported herein. Arsenic peaks with “fluvial sand/estuarine clay” units due to their higher adsorption capacity.

Khan et al.’s (2001a) reason for rejection of arsenic contribution by hinterland erosion

in Bangladesh was based on the non-proportional amount of opaque minerals present

within the alluvium that indicated arsenic was not derived from parent arsenic bearing

minerals. This theory was in turn examined at Stuarts Point. Common opaque minerals

include iron oxides, sulfides and heavy minerals such as ilmenite. Oxides and sulfides

have been identified in the aquifer and are discussed further in Chapter 6. Oxides are not

excellent indicators of hinterland contribution due to exposure and formation

(oxidation) upon the lowering of sea level in the Pleistocene. Concurrently, sulfides

(e.g. pyrite) are also questionable as hinterland indicators due to possible authigenic

precipitation of pyrite after sediment deposition. Heavy minerals have previously been

mined in close proximity to the Stuarts Point study area and their genesis is attributed to

wave action during the building of coastal sand barriers, with their original source

provenance terrigenous (Roy, 1999). The immediate source of the NSW east coast

heavy mineral sands seems to be the Mesozoic sediments of the Sydney, Clarence and

Moreton Basins (Thorp and McDonald, 2001), however sands on the mid-north coast of

NSW - close to Stuarts Point – are attributed to a New England Granite source (Wallis

and Oakes, 1990). Therefore, the use of heavy opaque minerals potentially supports

hinterland derived aquifer sediments for Stuarts Point. This has further been supported

by use of the indicator element, antimony.



The latter two studies cited by Smith et al. (2006) examine the experimental aspects of

arsenic adsorption onto clays, confirming clay ability to sequester arsenic, rather than

the potential for arsenic origination from transgressing seawater. As is demonstrated by

the arsenic concentration peaks in the “fluvial sand/estuarine clay” unit (Figure 5.23),

fine-grained sediments possess higher surface areas for elemental adsorption and

therefore this accounts for the increased arsenic concentrations in these units when

compared to the “fluvial sands” and “beach barrier deposits”. It is important to note that

arsenic concentrations peak in the (mixed) clayey sand units as opposed to the high

plasticity estuarine clays containing little or no sand content. This demonstrates the

importance of fluvially derived sediments correlating with the highest arsenic

concentrations and therefore supports the contribution of arsenic from erosion of the

hinterland.

Based on these findings, it seems prudent to dismiss the feasibility of arsenic within the

Stuarts Point aquifer being derived from the past inundation of the aquifer sediments by

seawater.

5.3.4 A note on arsenic and acid sulfate soils / pyrite

The source of arsenic deduced from preliminary studies by Smith et al. (2003; 2006)

was also proposed as being derived from ASS material. The literature review outlined

the common association of arsenic with pyrite via substitution mechanisms (Section

3.4.3.1), and as such, ASS material containing pyrite may be a potential source of

arsenic to the Stuarts Point groundwater system. Within the sandy aquifer investigated

herein, no distinct ASS formations were identified, however, adjacent clayey deposits in

the wetland show visible signs (surface scalds) of ASS formation and exposure. If

arsenic is sourced solely from theses ASS horizons, the distribution would be governed

by the presence of ASS. No such correlation has been observed herein.

The terminology describing the association of arsenic and pyrite can have double

meaning. If ASS horizons are disturbed and release arsenic into the groundwater

system, then pyrite can be considered as an arsenic source. If arsenic is present in the

groundwater system during in-situ precipitation of pyrite, then it may be considered a

current arsenic sink. Distinguishing between the two processes may not be possible.



Discrete, micron sized, clusters of pyrite were observed in the aquifer matrix (Chapter

6). For the purpose of this investigation, arsenian pyrite is identified as a current arsenic

sink. This will be supported and discussed further in subsequent chapters.

5.3.5 Direct Bedrock Contribution

Groundwater discharge from consolidated bedrock into the overlying alluvial aquifer

has the potential to transport dissolved arsenic into the Stuarts Point sandy aquifer.

Bedrock lithology and groundwater flow dynamics thus become an important part of the

arsenic source determination process.

Section 2.3.2 identified arsenic as a probable trace element associated with mineralised

bedrock. Drilling for these investigations encountered metamorphosed sediments in

close proximity to the granitoid intrusion of Mt Yarrahappinni (ML9; Figure 5.9). It is

thus hypothesised that in the northern portion of the aquifer, metamorphosed bedrock

directly below the sand aquifer, may contribute localised arsenic to the system. In order

to explore this hypothesis as a potential arsenic source, a hydraulic link needs to be

identified between the bedrock and the unconsolidated aquifer. Since very little physical

hydrogeology is known regarding groundwater movement in the aquifer, chemical

methods of investigation need to be utilised to establish any connection between

bedrock and unconsolidated sediments.

Recognised mineral deposits most likely to affect groundwater chemistry are the

molybdenum deposits and associated outward zoning of metal sulfides discussed in

Chapter 2. Bore ML9, which penetrated phyllite bedrock of the Nambucca (Pee Dee)

Beds exhibits a localised increase in both arsenic and molybdenum with increasing

depth in the aquifer (Figure 5.24).



Figure 5.24 As and Mo groundwater concentrations with depth in ML9 (left) and ML6 (right). Note the simultaneous increase in both elements as depth increases and weathered phyllite bedrock is penetrated at 20 mbgs in ML9. Contrasting, ML6 shows no distinct correlation between the two elements.

Arsenic and molybdenum are very strongly correlated (r2 = 0.84) in this bore, which

suggests the disseminated molybdenum mineralisation in the Nambucca (Pee Dee) Beds

close to the contact aureole may be contributing minor amounts of arsenic to the

groundwater in this portion of the aquifer. In contrast ML6, which penetrates shale

bedrock at distance from the intrusion, shows no increase between arsenic and

molybdenum with depth and no significant correlation between the two elements (r2 =

0.03). The shale bedrock of the Macleay Series is therefore not considered an important

source of aqueous arsenic to the aquifer.

In support of minor bedrock arsenic discharge, Smith et al. (2003) putatively linked

arsenic concentrations with mineralisation in bores installed wholly within the

Yarrahapinni fractured rock aquifer. No trace element correlations were presented to

support this hypothesis and analysis of their data set shows no observable correlation

with molybdenum. They noted that bores penetrating weathered gravel zones were

depleted in arsenic and concentrations reached a maximum of only 17 g L-1 in the

fractured rock aquifer. It is suggested that the fractured nature of the exposed aquifer is



conducive to geochemical reactions promoting arsenic mobilisation and therefore

dissolved arsenic contribution to the sandy aquifer through bedrock fractures is

expected to be minimal.

Given the fractured nature of the bedrock, however, it is possible that discharge of

bedrock groundwater into the sandy aquifer may occur through such structures. The

presence of a fault could allow groundwater to act as a conduit and discharge arsenic

rich water into the aquifer above. This theory would suggest the original source of

arsenic is in an aqueous form. Uncertainty regarding the presence and location of the

Kempsey Fault (Section 2.3.2) means potential groundwater discharge through this zone

cannot be specifically addressed. The aqueous groundwater data shows no indication of

groundwater mixing at depth in the locations sampled during this study.

The potentially localised arsenic bedrock source in the northern portion of the aquifer

(underlain by the Nambucca [Pee Dee] Beds) is not considered to be a dominating

origin of arsenic in the aquifer, given the widespread occurrence of arsenic in the

aquifer. Additionally, there is no reliable isotopic evidence (Section 2.3.3) to support

groundwater discharge from the bedrock, it is only hypothetically assumed.

5.4 A COMPARISON TO ARSENIC SOURCES IN OTHER NATURALLY ELEVATED ARSENIC ENVIRONMENTS

The reconstruction of the aquifer’s depositional history suggests the dominance of a

regionally derived arsenic source via the weathering and re-deposition of mineralised

deposits from the Macleay River hinterland. Several other studies (excluding

Bangladesh) have deduced arsenic from a hinterland source. Yu et al. (2000) conducted

a combined geochemical, mineralogical and sedimentological investigation to determine

the source and distribution of arsenic along the coastline of southwestern Taiwan. The

depositional environment of the arsenic contaminated aquifer is almost identical with

the geomorphic development at Stuarts Point; rivers deposited load resulting in the

formation of an advancing alluvial plain with lagoonal and barrier environments and the

current coastline processes observed today (Figure 5.25). Slate was determined to be

the source rock for the arsenic, originating in the Central Mountain range of Taiwan and



transported via two rivers (Hsilou Chi in the north and Kauping Chi in the south) and

deposited to the study area by long-shore currents.


Figure 5.25 Development of an arsenic affected aquifer in Taiwan. Slate debris, deposited via fluvial and long-shore processes, is the source of the arsenic (Yu et al., 2000).

A less common occurrence of arsenic in groundwater has been investigated in a

sandstone drinking water aquifer in Bavaria (Heinrichs and Udluft, 1999). The Keuper

Sandstone aquifer is characterised by shallow marine basin facies (estuarine, deltaic)

and terrestrial facies (fluvial) divided by a 20 km transitional zone between these two

facies types. Arsenic concentrations in groundwater were notably higher in the

terrestrial facies – up to 120 g L-1 - as compared to groundwaters from marine facies

which ranged from only < 1-4 g L-1 As. Similar patterns were observed with As

concentrations in the aquifer rock material. Terrestrial facies deposited under fluvial-

marine environments were deduced as source rocks for the arsenic in comparison to

marine beds that contain no arsenic. The degree of arsenic contamination within the

Keuper Sandstone aquifer is directly related to the distribution of these two facies.



5.5 CHAPTER SUMMARY

Examination of both anthropogenic and geogenic sources of arsenic to the Stuarts Point

aquifer has yielded five major arsenic source possibilities:

1. Arsenic derived from one or more anthropogenic activities;

2. Arsenic contributed via fluvial erosion of the NEFB hinterland and subsequent

deposition;

3. Arsenic captured and retained by aquifer sediments during past changes in sea level;

4. The oxidation and release of arsenian pyrite present in ASS material; and/or

5. Aqueous arsenic contributed via upwards vertical bedrock / fault conduit discharge.

The absence of distinct contaminant plumes emanating from potential anthropogenic

sources dismisses the dominance of a human-induced arsenic source. However, diffuse

application of fertilizers and/or pesticides may have contributed to minor amounts of

arsenic in the aquifer. The review conducted herein proposes a predominantly geogenic

source of arsenic in the aquifer, with strong support for a regionally derived arsenic

source via erosion of the mineralised hinterland and subsequent deposition of these

sediments to form the current aquifer sediments. The use of Sb as a hinterland indicator

element supports the designation of a regionally derived arsenic source for the Stuarts

Point aquifer. In addition, localised discharge of groundwater from arsenic-mineralised

bedrock may occur in the northern portion of the aquifer. The contribution of arsenic

from past sea level transgressions is considered unlikely as the dominant source of

arsenic at Stuarts Point.

Table 5.7 provides evidence both for, and against, each potential arsenic source at

Stuarts Point.



Table 5.7 Potential arsenic sources to the Stuarts Point aquifer and the supporting/opposing evidence associated with each suggested source.

PotentialArsenic Source

Supporting Evidence Opposing Evidence

Anthropogenic Potential widespread (diffuse source) application of arsenic

containing fertilizers, pesticides & cattle dipping solutions; potential cemetery & rubbish tip leachate

(point source)

Insufficient usage records & no well defined contaminant

traceability; lack of well defined plumes from point sources

RegionalGeology

Geomorphic model & statistical analyses indicate regional

geological material contributed to aquifer matrix; erosion of arsenic

rich mineral deposits within regional geology; correlation with antimony derived from regional

geology & therefore use of suitable hinterland derived

indicator element

Uncertainty regarding same source provenance for beach

barrier deposits & fluvial sands, meaning different origins of aquifer material may not be

conducive to a sole geological source of arsenic

Sea level Contribution

Arsenic accumulation in marine clay deposits; similar sea level

changes experienced in Bangladesh where sea level

contribution has been postulated as an arsenic source to their

alluvial aquifers

Bangladesh has an arsenic-rich volcanic ash deposit enriching the transgressing sea water –

Stuarts Point lacks such a deposit; accumulation in marine

sediments shows a positive correlation with arsenic and

marine clays, however, clays exhibit a large surface area

available for arsenic adsorption regardless of the arsenic origin

ASS association The presence of ASS horizons adjacent to the sandy aquifer and the known relationship between

pyrite and arsenic

No distinct ASS horizons were observed in the sandy portion of

the aquifer. No ‘plume’ of arsenic was observed adjacent to known ASS surface deposits.

BedrockConditions

Suggested mineralisation occurring in bedrock beneath

aquifer; aqueous arsenic increase with depth to bedrock (localised)

indicating potential arsenic bedrock source; potential fault

providing a conduit for groundwater flow into aquifer

above

Only localised occurrences of aqueous arsenic sources at

bedrock detected, which are not considered sufficient to affect arsenic distribution within the whole aquifer; no evidence of

high arsenic content in groundwaters in the bedrock


6 IDENTIFICATION OF CURRENT ARSENIC SINKS

The aquifer matrix was investigated in Chapter 5 to determine the likely source of

arsenic. This chapter explores the matrix in more detail, focussing on where the arsenic

is currently ‘being held’ in the aquifer (i.e., current arsenic sinks). Geochemically,

arsenic is influenced by many parameters (e.g. redox potential, pH, composition of the

surrounding solution) frequently making it mobile under natural aquifer conditions and

sensitive to slight changes in the geochemical environment. An understanding of the

chemical forms of solid phase arsenic present in the aquifer is critical to determining its

potential for transport in the groundwater system.

The chapter is structured as follows:

Section 6.1 builds on the sediment descriptions provided in Section 5.2 by

examining the aquifer mineralogy at the micron scale;

Section 6.2 deals with the sediment chemistry, particularly the occurrence of

arsenic, and any relationships observable between arsenic and other elements in

the matrix;

Section 6.3 examines the likely sinks of arsenic present in the aquifer, as

detailed in the literature review of Chapter 3. The following potential sinks are

examined:

o Adsorption to oxyhydroxides

o Association with pyrite

o Incorporation into clay minerals

o Sorption on organic matter

o Association with calcite

o Anion competition

For clarity and flow of discussion, results and interpretation are included in each

respective section.

Chapter 6 – Identification of Current Arsenic Sinks


6.1 AQUIFER MINERALOGY AND PHYSICAL PROPERTIES

The aquifer has been established as an unconsolidated deposit dominated by sand and

clay (Section 5.2). In summary, medium to coarse grained sand dominates the upper

part of the aquifer above fine sands of fluvial origin, estuarine clays and in-situ

weathered bedrock. Quartz dominates over minor feldspar and lithics. Gravel sized

fragments of lithic material (metamorphic in origin, sometimes up to 3cm in length) and

carbonates (often up to 80% in shell lenses) are sporadic throughout the aquifer and

representative of changes in depositional conditions. The sands form the main water-

bearing zone of the aquifer and are readily recharged by precipitation. Zones of water

table fluctuation are conducive to impregnation of organic material and form a

discontinuous lens of coffee rock (indurated sand) capable of reducing infiltration and

leading to the development of localised perched water tables.

Detailed analysis of the sediment matrix was undertaken on a Hitachi S4500 Field

Emission Scanning Electron Microscope (SEM) fitted with an ISIS Energy Dispersive

x-ray Spectrometer (EDS). Atomic ratios are semi-quantitative and normalised to 100%.

The full electron microprobe methodology is outlined in Chapter 4. Of the 36 collected

sediments, samples were divided into the geomorphic facies developed in Section 5.2.

After several samples from each facies had been examined under the SEM, a

photomicrograph of typical sediment grains observed within each facies was selected

and is shown in Figure 6.1.

The ‘beach barrier sand” unit has previously been visually identified as quartz rich

(Section 5.2). Individual grains semi-quantitatively analysed by the SEM are composed

of silica and oxygen, with a normalised Si/O ratio of 0.5 indicative of quartz (Figure

6.1a). No other elements were detected despite the irregular morphology observed on

the grain surface. These irregularities are thought to represent amorphous quartz

overgrowth structures.



Figure 6.1 SEM photomicrographs of typical grains observed for each geomorphic facies in the Stuarts Point aquifer. (a) beach barrier sand (b) fluvial sand (c) fluvial sand/estuarine clay and (d) bedrock clay. Spectral plots coincide with the white spectral boxes outlined in each photomicrograph.



The “fluvial sand” sample is an olive-grey medium-fine grained sand (Section 5.2),

composed of quartz with a small proportion of calcium most likely due to the presence

of shell material within this unit (Figure 6.1b). The presence of shell material (i.e.,

molluscs) within a fluvial unit is not unusual given the proximity of the aquifer to the

coast and the transitional nature of the aquifer within a small scale (i.e., on a metre

scale). Some overlapping of beach and fluvial sands is probable. Additionally, some

Australian molluscs (e.g. Alathyria jacksoni) live in freshwater. The small white

aggregates bound to the surface of the grain were also composed of quartz with minor

calcium.

The underlying “fluvial sand / estuarine clay” unit is finer grained, with particles

varying from approximately 10-300 m in the clayey sand (as observed under SEM).

This decrease in particle size has an influence on its major composition (Figure 6.1c).

Quartz is suspected to remain dominant, however a reduction in the Si/O ratio to 0.36

and the presence of other elements (K, Al, Mg, Na, Fe, Ca) indicates the potential

presence of clay minerals and oxides.

The heterogenous mix of mineral grains such as quartz, which is visible under standard

laboratory microscopes, with a variety of potential clay minerals and oxides, does not

allow identification of any specific minerals via atomic ratios reported by the SEM.

Surface structures can be seen on the larger particles even at a high magnification of

x100 in these sediments. Their semi-quantitative analysis indicates a similar

composition to the surface of the grain.

High levels of O and Si in conjunction with Al and K indicate the presence of a clay

mineral in the bedrock clay unit (Figure 6.1d). The atomic ratios for Si/O (0.2) and K/Al

(0.2) are close to illite composition. Bulk arsenic concentrations were below detection

limit for the SEM. The location of the arsenic peak is shown to indicate its absence.

6.2 HOW MUCH ARSENIC IS IN THE AQUIFER MATRIX ?

Arsenic-rich groundwaters are not frequently associated with arsenic-rich host matrices.

It is common for arsenic concentrations in sediments to be in the range of normal crustal



abundance; 1-20 mg kg-1 As. Such is the case for Stuarts Point. Arsenic concentrations

in the aquifer matrix range from 1.4 to 14.0 mg kg-1 with a mean concentration of 6.3

mg kg-1 (n = 36).

Similar concentrations have been observed in Bangladeshi sediments where arsenic

concentrations were reported to range from 0.4 to 10.3 mg kg-1 (n = 21) for three

spatially distant districts (BGS and DPHE, 2001). This relationship between low solid

phase arsenic and high aqueous arsenic concentrations was noted by Smedley and

Kinniburgh (2002) who suggested it appeared to be the rule rather than the exception

and may be a characteristic responsible for previously delaying the identification of high

arsenic groundwaters in many regions.

The use of mean elemental concentrations in a heterogeneous aquifer, however, can bias

interpretation. Clays can sequester considerably higher concentrations of trace elements

when compared to the values reported for sandstones. This variation can be expected

between the different geomorphic facies in the Stuarts Point aquifer.

To determine element anomalies for the various geomorphic facies identified at Stuarts

Point, sediment enrichment factors (EF’s) were calculated. Comparison of sediment

element concentrations to similar sedimentary environments can indicate possible

enrichment of specific metals and trace elements relative to their sediment lithology. A

common method of assessing this phenomenon is via the calculation of EF’s.

Table 6.1 lists the 22 trace elements analysed herein and their reported means in the

Stuarts Point aquifer sediments. Background mean element concentrations for various

rock and sediment types have been provided for indicative comparison. These

background concentrations highlight potential element anomalies in the Stuarts Point

sediments, as compared to natural element concentrations for various aquifer matrices.

For instance, mean Sb concentrations at Stuarts Point are 6 mg kg-1 while all other

average values are reported at or below 1.5 mg kg-1.



Table 6.1 Elemental mean and ranges (mg kg-1) reported herein for the Stuarts Point aquifer sediments and their comparison to natural background concentrations for different matrices (full Stuarts Point data set provided in Appendix B1).

Stuarts Point Sediments Mean Concentrations Reported in Literature Element

Mean Range Shale1 Sandstone2 Granite1 Clay1

As 6.30 1.40 - 14.0 13.0 2.00 1.50 13.0Ba 178 0.00 - 849 580 20.0 840 2300Ce 35.2 16.1 - 108 59.0 6.00 92.0 345Cd 0.20 0.10 - 3.70 0.30 0.05 0.13 0.42Co 32.8 0.00 - 113 19.0 0.30 1.00 74.0Cr 22.3 0.00 - 108 11.0 35.0 4.10 90.0Cu 13.3 5.60 - 64.6 45.0 2.00 10.0 250Ga 5.80 2.10 - 18.8 19.0 12.0 17.0 20.0Mo 1.00 0.00 - 5.50 2.60 0.20 1.30 27.0Nb 3.40 0.90 - 10.5 11.0 0.001 21.0 14.0Ni 7.70 0.00 - 36.4 68.0 2.00 4.50 225Pb 7.30 2.60 - 25.7 20.0 12.0 19.0 80.0Rb 36.1 7.50 - 106 140 60.0 170 110Sb 6.00 1.30 - 11.4 1.50 0.40 0.20 1.00Sn 0.00 0.00 - 0.00 6.00 0.001 3.00 1.50Sr 76.7 11.2 - 227 300 20.0 100 180Th 3.20 0.00 - 13.5 12.0 5.50 17.0 7.00U 1.80 0.00 - 5.00 3.70 2.00 3.00 1.30V 48.6 10.6 - 216 130 20.0 44.0 120Y 8.50 0.00 - 55.1 26.0 10.0 40.0 90.0Zn 21.3 4.50 - 114 95.0 16.0 39.0 165Zr 94.5 32.3 - 244 160 220 175 150

1 Turekian and Wedepohl (1961). 2 Drever (1997) and Tucker (1991)

Loska et al. (2004) describe the EF as the comparison of a tested element against a

reference one according to the following equation:

EF = (cn [sample] / cref [sample]) (6.1)

(Bn [background] / Bref [background])

where cn [sample] is the concentration of the examined element in the examined

environment, cref [sample] is the concentration of the reference element in the examined

environment, Bn [background] is the concentration of the examined element in the

reference environment and Bref [background] is the concentration of the reference

element in the reference environment. A reference element is a conservative element

that is chosen specifically for a site based on its correlation with elements of concern

and its low potential to be present due to anthropogenic activities. Aluminium was

selected as the reference element for this study because it exhibited the highest r2



correlation with the majority of trace elements (Appendix B7). Similarly, Liaghati et al.

(2003) chose aluminium as the reference element for a coastal plain catchment in

nearby Queensland, Australia. For comparison of this environment to the Stuarts Point

coastal environment it is advantageous if the reference element is the same.

Of particular interest to the Stuarts Point study, is the proposed derivation of aquifer

sediments from regionally eroded geology abundant with various types of

mineralisation. To ascertain whether the Stuarts Point sediments are unusually elevated

in some trace elements, comparison to a similar coastal or geomorphological unit is

required. EF’s were thus calculated using mean elemental concentrations of average

sandstones (for the “beach barrier sands” and “fluvial sands”), mean elemental

concentrations of average clays (for the “bedrock clays”) and mean concentrations of

pre-industrial estuarine sediments reported for a coastal environment in Queensland

(Preda and Cox, 2002) for the “fluvial sand / estuarine clays” at Stuarts Point (Table

6.2).

The comparison to Queensland estuarine clays is particularly important. One of the

initial arsenic source theories proposed by previous investigators (Smith, 2006; Section

2.4) is that arsenic is contained within the Stuarts Point aquifer due to inundation of the

aquifer by sea level rise in the Quaternary. If this source theory is correct, then the entire

east coast of Australia (which was subjected to the same sea level movements) may be

at risk of elevated arsenic. Additionally, the first paragraph of this thesis (Section 1.1)

explains that the discovery of arsenic in an aquifer south of Stuarts Point may lead to

concern for arsenic to be present in other coastal sand aquifers of eastern Australia. It

was these results that led to the inception of this project. Therefore, Queensland

provides an excellent example of a background environment, given its coastal location

and the same sea level movements as endured at Stuarts Point.

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2.1

na1.

80.

3M

L7/1

818

.60.

6na

na1.

1na

nana

0.8

1.2

na0.

53.

72.

64.

21.

6na

1.4

0.2

ML7

/22.

513

.50.

7na

na1.

8na

nana

0.7

1.0

na0.

84.

33.

52.

31.

3na

1.7

0.4

ML7

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20.0

1.8

nana

1.5

nana

na0.

90.

9na

0.8

4.4

3.7

5.9

1.6

na2.

20.

3M

L8/1

96.

90.

2na

na0.

4na

nana

0.5

0.3

na0.

62.

32.

23.

40.

8na

1.1

0.4

ML8

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1.4

0.1

nana

0.2

nana

na0.

50.

4na

0.6

2.2

2.3

4.0

0.9

na1.

10.

4M

L9/1

9.1

4.5

0.1

nana

0.2

nana

na0.

50.

2na

0.5

1.7

2.0

10.7

0.7

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00.

4M

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0.6

6.5

0.1

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0.3

nana

na0.

70.

7na

0.7

2.1

2.4

34.4

1.0

na1.

40.

4M

L10/

13.3

8.5

0.2

nana

0.3

nana

na0.

50.

2na

0.5

1.5

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53.8

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4na

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1.6

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11.3

0.9

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0.8

nana

na0.

60.

5na

0.7

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2.9

10.2

1.2

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22.3

4.3

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0.2

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1.7

2.3

4.0

1.1

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30.

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L10/

25.3

7.3

0.1

nana

0.2

nana

na0.

60.

3na

0.5

2.4

2.7

16.6

1.4

na1.

80.

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If arsenic is in fact derived from the sea, then both the Queensland sediments and the

Stuarts Point sediments may have similar arsenic compositions. This was not found to

be the case. Arsenic was higher in the Stuarts Point sediments. Other trace elements

provided similar results. These elevated trace elements can be found in the hinterland of

the Macleay River, thus providing evidence that the arsenic at Stuarts Point is derived

from the hinterland, not sea level transgressions during the Quaternary.

Another possible reference environment could have been selected for the estuarine

clays; the Stuarts Point bedrock clays. However, the bedrock clays are influenced by

mineralisation in the bedrock below. Therefore, they would bias the results. For the

most part (except the few bedrock clay samples), the aquifer clays are not derived from

weathering of the bedrock. The clays are the result of estuarine processes. The

Queensland background material were once again excellent references due to their

estuarine nature – deposited in a manner most similar to the estuarine clays found at

Stuarts Point. It is for these reasons that the bedrock clays would be deemed

inappropriate for comparison to the estuarine clays within the aquifer.

According to Sutherland’s (2000) five categories based on the enrichment factor (Table

6.3) the following enrichments occur within the Stuarts Point aquifer matrix:

1. Moderately to extremely highly enriched in antimony in all (bar one) sediment

samples. This uniform enrichment (across all depth intervals, rather than surface

samples only) dismisses the possibility of antimony as an anthropogenic

contaminant and supports the proposal of antimony derivation from regional erosion

of stibnite deposits in the upper reaches of the Macleay River catchment;

2. Extremely highly enriched in cobalt in all sands contained within the aquifer. The Bn

[background] concentration for Co in sandstones has been listed as 0.3 mg kg-1

which may be considered low, however even when this value is raised to a crustal

abundance of 20 mg kg-1 (Bowen, 1979) it remains significantly enriched,

particularly in the “fluvial sands”. Cobalt is commonly associated with Ag, Ni, Pb,

Cu and Fe ores, but most abundant in sulfide ores (Adriano, 1986). Therefore cobalt



can also be considered an indicator of hinterland erosion and depositional processes;

and

3. Enriched to various degrees in As, Mo, Cu, Ce, Ni and Sr throughout the aquifer.

Arsenic, Mo and Cu mineralisation all occur in the NEFB, while the majority of Ni

minerals are sulfides or arsenides. Strontium can be associated with granite or shell

material and Ce may be the product of monazite (i.e., heavy mineral sands) present

in the aquifer.

Table 6.3 EF Categories devised by Sutherland (2000). EF Value EF Category

< 2 Deficiency to minimal enrichment 2 –5 Moderate enrichment

5 – 20 Significant enrichment 20 – 40 Very high enrichment

> 40 Extremely high enrichment

The enrichment of arsenic is particularly noted for all “fluvial sand” samples, in

comparison to no enrichment observed in the “fluvial sand / estuarine clay” unit1. This

may be an indication of arsenic enrichment via fluvial processes. Copper, Co and Ba are

enriched in a similar way, suggesting possible element relationships or the same source

provenance.

The enrichment of elements indicative of mineralisation in both the “beach barrier

sands” and the “fluvial sands” suggests these sediments are derived from a regional

geologic source in comparison to local anthropogenic inputs. Thus, the use of EF’s

supports the derivation of arsenic and other trace elements (particularly Sb) from

regional geology.

1 This is of course a direct response to the higher element concentration used for the reference environment value but which takes into account the higher surface area for arsenic sorption for clays, in comparison to sands.



6.2.1 Arsenic-bearing Minerals

A selection (n = 9) of quartz-rich sand, coffee rock and estuarine clay sediments was

analysed by XRD to determine mineralogical composition. Most results produced

quartz rich XRD spectra. Calcite was identified in sand samples where shell specimens

were recovered and minor feldspar was detected in some samples. Illite (and minor

kaolinite) was identified as the clay type present in the aquifer but quartz remains

dominant overall.

Figure 6.2 shows the XRD spectrum for a fluvial sand/estuarine clay in ML7. As can be

seen by the assignment of mineral peaks on the spectrum, quartz dominates the

sediment composition. Minor amounts of feldspar, illite and kaolinite are also identified.

XRD indicated the presence of minor amounts of pyrite in samples with high clay

content such as Figure 6.2. Smaller crystals were probably not identified due to their

discrete size and potentially amorphous form. No arsenic-bearing minerals, such as

arsenopyrite, were identified.

Figure 6.2 XRD spectrum for a fluvial sand/estuarine clay at ML7/27. Quartz dominates with minor amounts of illite, kaolinite, feldspar and pyrite identified.



6.2.2 Are there any distinctive physical or geochemical correlations ?

Arsenic geochemistry is complex. Often concentrations in solid phase can be

representative of original arsenic sources, and those that have been redistributed through

various sorption, precipitation and dissolution mechanisms.

For example, arsenian pyrite may oxidise and release arsenic into solution, with iron

oxide by-products scavenging this released arsenic and retaining it in solid form.

Changes to iron oxide morphology, solution pH and Eh, or addition of competing ions

can release arsenic into solution once again. The slow redox transformation of arsenic,

aquifer heterogeneity and anthropogenic use of the aquifer may leave arsenic re-

distributed in different forms throughout the aquifer. Therefore, the correlation between

arsenic and other element concentrations in the solid phase may only be indicative of

potential geochemical sinks in different parts of the aquifer. By statistically clustering

aquifer sediments into more homogenous geomorphic units, aquifer heterogeneity is

somewhat addressed. Examination of arsenic occurrence in each of these units can then

be independently assessed.

Pearsons correlations (Appendix B7) were calculated to evaluate arsenic geochemical

relationships in each facies (Table 6.4). For the purpose of discussion, the following

terminology applies herein:

r2 = 0.91 to 0.99 extremely strong correlation

r2 = 0.81 to 0.90 very strong correlation

r2 = 0.61 to 0.80 strong correlation

r2 = 0.41 to 0.60 moderate correlation

r2 = 0.31 to 0.40 weak correlation



Table 6.4 Pearsons correlations between solid phase arsenic and trace/major element concentrations reported for the Stuarts Point geomorphic facies. Extremely strong correlations are highlighted in black, very strong correlations in grey, and strong correlations are bolded. Additional X-Y plots for selected correlations with arsenic are located in Appendix B9.

Barrier Sands

Fluvial Sands

Fluvial Sands / Estuarine Clay

BedrockClay

Sb -0.06 -0.36 -0.52 -0.76Mo -0.24 0.30 0.22 0.60Nb -0.04 0.04 0.70 -0.97Zr 0.25 0.19 0.74 -0.90Y 0.02 0.68 0.44 -0.50Sr -0.21 0.58 0.15 -0.77U -0.42 0.12 -0.16 -0.48

Rb 0.02 0.77 0.64 -0.96Th 0.18 0.66 0.32 -0.92Pb 0.35 0.73 0.11 0.38Ga 0.65 0.84 0.54 -0.99Zn -0.53 0.76 0.57 0.00Cu -0.48 -0.35 0.21 -0.01Ni -0.44 0.74 -0.43 -0.45Co -0.37 -0.72 0.27 0.29Cr 0.01 0.44 0.42 0.78Ce 0.12 -0.39 0.59 0.46V 0.48 0.68 0.62 -0.76

Ba 0.24 0.67 0.59 -0.93LOI 0.09 0.45 0.30 -0.81SiO2 -0.24 -0.73 -0.57 0.95TiO2 0.34 -0.23 0.75 0.15Al2O3 0.74 0.65 0.73 -0.98Fe2O3 0.70 0.93 0.01 0.68MnO -0.12 0.01 0.16 -0.06MgO -0.29 0.84 0.14 -0.66CaO -0.15 0.53 0.26 -0.09Na2O 0.25 0.76 0.22 -0.57K2O 0.42 0.70 0.14 -0.98P2O5 0.08 0.54 0.32 0.90SO3 0.08 0.78 0.38 -0.09

Correlation with Arsenic



The identification of average arsenic concentrations (i.e., low background

concentrations) in the aquifer sediments at Stuarts Point supports the hypothesis that

arsenic is naturally derived rather than anthropogenically induced. Frequently associated

with naturally occurring arsenic are the elements iron, aluminium and manganese

(representing silicates, oxyhydroxides or sulfides); various trace elements derived from

ore deposits (molybdenum and antimony have previously been shown to be allied with

arsenic regionally); species of sulfur (commonly delineating a pyrite association); loss

on ignition (LOI) indicating potential arsenic association with organic matter; fine grain

size correlation with elevated arsenic concentrations indicating silts and clays

potentially immobilising arsenic; and phosphate or bicarbonate ions dueling for the

same aquifer sorption sites as arsenic. Thus, a correlation between arsenic and one or

more of these elements or parameters could identify the presence of particular solid

phase arsenic sinks.

Examining these correlations in order of oldest sediments to youngest, the in-situ

“bedrock clays” exhibit the greatest element correlation with arsenic. Niobium and Th

are often hosted by Ti-oxides (Lopez et al., 2005); Rb and Ba have been associated with

sorption on clays (Eylem et al., 1990; Kleven and Alstad, 1996; Lopez et al., 2005); Ga

is generally associated with clays of granite source (Burton et al., 1959) while LOI,

Al2O3 and K2O5 may be direct representations of clay mineral occurrence. Thus, the

high negative r2 values reported with arsenic for Nb, Rb, Th, Ga, Ba, LOI, Al2O3 and

K2O5 are suggested to indicate a non-association with a potassium clay likely to be illite.

In other words, arsenic is not associated with the weathered bedrock clays. The

remaining correlation with arsenic for the “bedrock clays” are extremely strong for SiO2

(r2 = 0.95) and very strong for P2O5 (r2 = 0.90). This suggests a relationship between

arsenic and the quartz fraction perhaps partially coated with iron oxide (r2 = 0.68) which

is also sorbing phosphorous.

The “fluvial sand / estuarine clays” exhibit a strong positive correlation between arsenic

and Nb, Zr, Rb, V, TiO2 and Al2O3 which is in contrast to that previously observed for

the bedrock clays. Plausible explanations could include the presence of a different clay

mineral with different sorptive properties, however this is unlikely as XRD has

identified illite in both geomorphic units. Alternatively, the heterogenous mix of sands



and clays in this unit may influence the statistical results. It is interesting to note that

arsenic and iron are not correlated (r2 = 0.01) for this geomorphic unit.

A better correlation with iron is observed for the “fluvial sands”. In contrast to the

absence of an iron oxide and arsenic correlation in the underlying sediments, an

extremely strong positive (r2 = 0.93) correlation exists for As and iron in the sands of

purely fluvial origin. In addition, the strong positive correlations with clay indicative

elements and SO3, with a strong negative SiO2 (r2 = -0.73) correlation suggests that iron

oxide, clay and sulfur mineral coatings may be dominating the fluvially transported

quartz grains which are sorbing arsenic to their outer edges.

Arsenic association with elements in the “beach barrier sands” results in strong

correlations for both Al2O3 (r2 = 0.74) and Fe2O3 (r2 = 0.70). Clay dominance or sulfur

association is not indicated therefore suggesting arsenic association with oxide grain

coatings in this highly re-worked homogenous sand unit. Alternatively, association with

coffee rock may explain the aluminium association in the absence of other clay

indicators. A correlation with LOI is not expected in the “beach barrier sands” as coffee

rock is sporadic and localised and generally organic matter contents are low throughout

the unit.

A moderate negative correlation (r2 = -0.58) has previously been proposed between

arsenic and antimony for all sediments (bulk statistical result) regardless of their

geomorphic origin. This correlation is proposed to be indicative of arsenic and antimony

being derived from the same source with arsenic being mobilised after deposition while

antimony is retained in the solid phase. When each geomorphic unit is assessed

independently the correlation between the two elements is almost non-existent in the

surface sediments (r2 = -0.06 for the “beach barrier sands”) yet increases with depth,

reporting a strong negative (r2 = -0.76) correlation in the “bedrock clay” unit. The

barrier sands deposited in the top portion of the aquifer are depositionally youngest, yet

have been derived from the NEFB, deposited offshore and re-distributed onshore as

barrier sands. They were most likely eroded from the NEFB earlier than the underlying

fluvial sediments. Their extensive re-working has altered the original Sb-As signature

resulting in low statistical correlation. This signature remains recognisable in the



underlying sediments that have not been extensively re-worked, particularly the bedrock

clays which may contribute Sb from in-situ weathering of the granitoid influenced

bedrock. Arsenic-rich pyrites in a mineralised granitoid intrusion investigated by

Zacharias et al. (2004) also found the pyrites and associated arsenopyrites to be enriched

in antimony. The metasediments surrounding the contact aureole of Mt Yarrahapinni

have been described as pyritic (Leitch, 1974).

The association of arsenic with molybdenum was discussed in relation to bedrock

contribution of arsenic to the aquifer. Molybdenum is present in disseminated form in

the northern part of the bedrock aquifer where arsenic has also been associated with

these mineral deposits. This association of As and Mo with depth is confirmed via the

moderate positive correlation (r2 = 0.60) observed between the two elements in the

“bedrock clays” compared to no statistically significant correlation observed for the

other geomorphic units.

Manganese oxides exhibit a high affinity for arsenic sorption (although to a lesser extent

than iron oxides) and have been suggested as arsenic sorbents for many

hydrogeochemical studies (Mucci et al., 2003; Chaillou et al., 2003). Coupled with the

low concentrations of Mn in the Stuarts Point aquifer matrix (up to 0.04%), no

statistical correlation was observed with arsenic in the solid phase thereby reducing the

likelihood of Mn oxides as a potential arsenic sink.

Arsenic inclusion into the calcite structure was documented in the literature review. A

moderate positive (r2 = 0.53) correlation between As and Ca was observed in the shell

dominated “fluvial sand” unit potentially suggesting an arsenic-calcite association at

localised depths. This correlation is similar to that between arsenic and phosphorous for

the same geomorphic unit indicating phosphate inclusion into the calcite lattice as a

possible simultaneous geochemical process. Dual anion interaction is also expected

between phosphate and arsenic in the “bedrock clays” where sorption is dominated by

iron oxides. This is supported by the very strong correlation (r2 = 0.90) between As and

P2O5 in this unit.



Given that smaller grains (silts and clays) have greater surface areas for sorption

capacity, it can be expected that arsenic concentrations will be higher in sediments with

more clay content. This has previously been shown in vertical plots showing arsenic

concentrations increasing in the “fluvial sand / estuarine clay” units (Figure 5.23), when

compared to lower arsenic concentrations reported in the more sandier units of the

aquifer. Figure 6.3 shows that the “fluvial sand / estuarine clays” and “bedrock clays”

have the highest arsenic concentrations in solid phase in comparison to the lowest

concentrations observed in the “beach barrier sands”, supporting the association

between arsenic and finer grained sediments.

Figure 6.3 As versus fine fraction component (silts and clays) in the aquifer sediments. The clayey units report the highest As concentrations yet no good correlation exists to support As association with just the fine fraction.

Dudas (1987) found that arsenic was distributed over all size ranges but about four

times higher in the smaller size fractions. He attributed these results to arsenic sorption

on surface sites which were more abundant in the smaller size fractions. He also found,



“The broad partitioning of As among size separates suggests the element was

adsorbed onto iron oxide-coated mineral surfaces rather than occurring in a

discrete mineral or as an inorganic precipitate.” (pg 327)

The lack of correlation throughout the Stuarts Point aquifer between arsenic and the

amount of fine grained material can be explained by grain coatings present in all

geomorphic units contributing to arsenic distribution, rather than the singular

dominance of arsenic sorption onto clays.

These statistical correlations indicate numerous controls on arsenic distribution in the

solid phase and therefore support the assumption that arsenic immobilisation factors are

geochemically complex. Further solid phase analysis was thus conducted to assess each

of the main arsenic sinks prior to analysis of aqueous geochemical conditions.

6.3 EXAMINATION OF KNOWN ARSENIC SCAVENGING MATERIALS IN RELATION TO THE STUARTS POINT AQUIFER MATRIX

The statistical element correlations indicate certain geochemical controls on arsenic

distribution, however their likelihood of arsenic retardation cannot be confirmed until

their presence in the aquifer is established. The literature review outlined various

experimental studies on arsenic sorption phenomena for different minerals. Each of

these potential sinks, and their likelihood of retarding arsenic at Stuarts Point, is

discussed below.

6.3.1 Adsorption to Oxyhydroxides

Smedley and Kinniburgh (2002) state;

“Fe oxides are probably the most important adsorbents in sandy aquifers because

of their greater abundance and the strong binding affinity” (page 534).

The good statistical correlation between arsenic and iron in units of the aquifer solid

phase indicates arsenic sorption may be occurring onto iron oxides. Two important

questions supporting this potential theory remain unanswered:

1. Are iron oxides present in the aquifer?



2. Do aqueous geochemical conditions support sorption of arsenic to iron oxides?

(Chapter 7)

A common method for assessing arsenic fractionation in sediments is via selective

extraction methods where specific fractions of the soil are targeted by chemical

extractants and arsenic association with each fraction is quantified. There are distinct

advantages and disadvantages in using these analytical methods. The advantages

include obtaining an indication on the relative mobility of elements in comparison to a

total elemental analysis that does not indicate the potential for element transport through

the matrix. The disadvantages lie mostly with the complexities involved in the chemical

extractants used and the analytical methods employed. For example, errors can be

introduced in the sample drying stage (Farrah and Pickering, 1993); due to incomplete

dissolution of a targeted fraction (La Force and Fendorf, 2000; Gleyzes et al., 2002b);

redistribution of trace elements due to precipitation with extracting chemicals (Shuman,

1982); not to mention contamination of extraction solutions, variabilities in reaction

times and lack of quality control (Gleyzes et al., 2002a). Given these disadvantages,

some authors question the validity of using selective extraction techniques while others

recognise their shortfalls and assess the results accordingly.

There are numerous studies utilising selective extractions, each one different depending

on the nature of the matrix and the particular phases being targeted. Often they are

employed in studies on contaminated sediments and sludges (Carlsson et al., 2002; Bird

et al., 2003; La Force et al., 2000) where contaminants are present in high

concentrations and results can be verified to some extent via SEM analyses confirming

the extractant solution has removed the targeted fractions. Variations to sequential

extractions need to be made when the elements of concern are anionic rather than

cationic because sorption behaviour differs with pH. In addition, arsenic is sensitive to

changes in redox and some extractant chemicals can induce oxidation, which may

influence arsenic associations in the solid phase. Subsequently, several specific schemes

have been proposed especially for arsenic (Keon et al., 2001; van Herreweghe et al.,

2003; Wenzel et al., 2001).



Another problem encountered with selective extractions is their applicability to natural

sediments. As mentioned, arsenic concentrations in sediments can be low in areas where

high arsenic groundwaters exist. Dudas (1987) used extractant chemicals commonly

performed on arsenic contaminated sediments but produced no effective results in his

natural sediments. This may be a major drawback for assessing arsenic fractionation in

sediments of naturally high arsenic groundwaters.

Sequential extractants were trialled herein in the hope they might produce reasonable

results (O’Shea and Jankowski, 2003). An arsenic specific scheme was employed

utilising extractant chemicals from several studies, their advantages and disadvantages

noted in Chapter 4. Five specific pools were targeted:

Exchangeable – arsenic sorbed to mineral surfaces that can be released by changes

in solution pH or competition with phosphate;

Carbonate bound – arsenic incorporated into the lattice sites of calcite which are

acido-soluble;

Reducible – arsenic associated with amorphous and/or crystalline oxyhydroxides of

iron and aluminium;

Oxidisable – arsenic associated with organic matter and partial sulfide oxidation;

and

Residual – arsenic bound in the residual mineral phase such as in silicate lattice sites

and therefore not prone to mobilisation.

As found by Dudas (1987), the naturally low levels of arsenic in the Stuarts Point

matrix were not sufficient for chemical extraction, thereby reporting no detectable

arsenic in the extracting solutions. This is due in part to the large dilution factors applied

to enable the chemical extractants to be analysed via ICP-MS; and limitations with

laboratory equipment being small enough to fit into the available water bath and

centrifuge, which subsequently led to small sediment to aqueous extraction ratios. In

addition, precipitation reactions between phases is strongly suspected given that

phosphate, which had previously been used as an extractant chemical and was therefore

present at significant concentrations, decreased to levels below detection when the

reducible phase was targeted, indicating redistribution as an insoluble phase. Given that



phosphate has similar chemical properties to arsenate, similar redistribution may have

occurred to effect the arsenic results.

Despite the unfortunate fact that the method used to quantify the principal target

(arsenic) did not produce reasonable results, the findings of the trial were not discarded.

The amount of iron released during the reducible phase was used to infer what

percentage of total iron might be present as iron oxide (Appendix B5). A similar method

was used to determine aluminium oxide presence (Table 6.5).

Table 6.5 Indicative amount of iron and aluminium present as oxides in the Stuarts Point matrix.

Trial Sample for Reducible Phase

% Total Fe as iron oxide

Total iron(% XRF)

% Total Al as

aluminiumoxide

Totalaluminium

(% XRF)

Farm Soil 0.00 0.20 0.59 0.25Coffee Rock 8.34 0.60 0.78 1.24

Barrier Sand (ML9/5.6) 0.00 0.23 0.04 0.90Fluvial Sand / Estuarine

Clay (ML7/18) 1.49 1.10 0.07 3.64

Bedrock Clay (ML10/26.8)

0.20 9.05 0.03 6.00

Ferrihydrite, with its high surface area and adsorptive capacity, can sorb ions from

solution even when it is present in very small concentrations in sediments (Childs,

1992). Taking this into account, reasonable amounts of iron oxide are indicated to be

present in some parts of the aquifer, particularly associated with coffee rock.

Aluminium oxides have less affinity for arsenic sorption, but when present in significant

quantities can contribute to its retardation. The extractant chemical used for the

reducible phase may not be the most effective for aluminium oxides (Gleyzes et al.,

2002a). In combination with the low total aluminium and lack of strong correlation with

arsenic, the minor amount of Al suspected to be attributable to aluminium oxides are not

expected to be a dominant arsenic sink within the aquifer.

As an additional investigative technique to enable further support for iron oxide

presence, particle surfaces in the sediment matrix were observed using SEM. Arsenic

was not expected to be detected during SEM analysis given the highest reported As



concentration (14 mg kg-1) by XRF was well below the As detection limit for the SEM

(approximately 50 mg kg-1). Thus, the aim of using the SEM was to identify

morphological irregularities at the particle surface and semi-quantitatively analyse the

particle surface composition. Once individual particles and/or mineral precipitates have

been semi-quantitated, their ability to scavenge and/or contribute arsenic to the system

can be critically assessed2.

Minor amounts of iron were detected by SEM in samples throughout the aquifer,

particularly in the coffee rock and clayey samples. Atomic ratio calculations generally

did not distinguish discrete iron oxide phases, more likely indicating iron association

with the clay mineral structure and/or minor oxide presence. However, the bedrock at

the base of ML10 was notably different to other samples. Arsenic concentration

reported by XRF was 14 mg kg-1, the highest reported for all samples analysed. This

sample exhibited a distinctive bright orange colour and contained lithic fragments up to

3 mm in size. It is assumed to represent weathered bedrock at the base of the sandy

aquifer containing some fluvially deposited river gravels. It is hypothesised that the

ubiquitous olive grey clays existing in other places throughout the aquifer are under

reducing conditions, and that ML10/26.8 has undergone oxidation when the sea

regressed during the Pleistocene.

Upon first examination the clay component of this sample was logged as a potential

oxidised iron sulfate such as jarosite. The arsenic was thought to be present as a re-

distributed phase associated with jarosite or other oxidation products, which is what

gave this sample its distinctive colour and elevated arsenic concentrations. SEM was

expected to identify jarosite and confirm this hypothesis. Detailed analysis of the

sample was undertaken (Figure 6.4). The position of the arsenic peak is labelled on

each spectra to indicate its potential presence. The first point semi-quantitated was

chosen to represent the bulk particle surface (i.e. the smooth, unaltered grain surface).

2 The semi-quantitative SEM-EDS analysis, coupled with atomic ratios influenced by variable numbers of hydrated ions, enforces the inexact nature of the spectral analysis and the importance of assessing approximate element/atomic percentages only.



Figure 6.4 Detailed analysis of particles within oxidised bedrock clay (ML10/26.8) with some fluvial influence. Potential iron oxide minerals, as proposed by atomic ratios, are (a) goethite (b) ferrihydrite aggregate on goethite surface (c) amorphous iron oxide on goethite surface.

The spectrum (Figure 6.4a) suggests that the area outlined by the analysis contained

within the spectra box is rich in iron. Iron was reported at >50% with oxygen

contributing >30% element weight. The iron to oxygen atomic ratio is 0.50 indicating

this particle may contain goethite (FeOOH) since their atomic ratios are close to 0.50.

Arsenic bearing sulfides such as arsenopyrite, exposed to oxidising conditions (i.e.,

weathering processes) produce fine grained, secondary Fe(III) oxides/hydroxides



(goethite, ferrihydrite) and sulfates (jarosite) (Bowell, 1994). These ores are potentially

located in the bedrock below the alluvial aquifer, in addition to natural erosion of

mineral deposits in the upper reaches of the catchment contributing oxidised products

via fluvial deposition. No sulfide was observed by SEM, therefore dismissing the initial

description of the sediment as an oxidised sulfate. The SEM results suggest it is

predominantly an iron oxide.

Small cloudy-looking aggregates (up to 8 m but many much smaller) are present on

the goethite surface (Figure 6.4b). This aggregate was probed in order to determine if its

composition was similar or different to the particle surface examined in the previous

analysis. The results showed a decrease in the Fe/O ratio from 0.5 (previously) to 0.42

– the ratio indicative of ferrihydrite (Fe5HO8.4H2O). Schwertmann (1988) produced

TEM photographs indicating ferrihydrites were spheroidal with diameters ranging from

2-7 nm. The spheroidal shape and varying size of the aggregates observed herein

support the atomic ratios indicative of ferrihydrite.

Ferrihydrite is a less crystalline iron oxide than goethite. Iron(III) oxides recently

precipitated after oxidation of Fe(II) are poorly crystalline (Willett and Higgins, 1978)

suggesting that the goethite surface may be prone to oxidation. Alternatively,

transformation of ferrihydrite to goethite can occur after long periods of weathering

(Childs, 1992). A third, smaller aggregate, was also analysed on this goethite particle

(Figure 6.4c) to confirm it wasn’t a different mineralogical occurrence to the previous

aggregate. Its composition was intermediate between both points previously analysed.

The Fe/O ratio is 0.46 which is also intermediary between goethite and ferrihydrite.

Minor contributions of silica are expected as natural ferrihydrites can contain up to 9%

Si (Childs, 1992). Common Si/Fe ratios proposed by Childs range from 0.1 - 0.35.

Investigations reporting higher Si concentrations on less pure ferrihydrite samples

(Childs et al., 1982) were predicted to not all be attributed to Si contained within

ferrihydrite. The three iron oxides exhibit ratios of 0.35, 0.41 and 0.32 respectively.

These ratios account for Si inclusion in the oxide structure, perhaps with minor clay

mineral association for the ferrihydrite which also contains potassium. Ferrihydrite can

coat and bind kaolinite particles (Jones and Saleh, 1987) which may account for the



minor elemental impurities detected. The presence of aluminium in each iron oxide

may be due to isomorphic substitution for Fe as suggested for goethite and hematite by

Zhao et al. (1994).

A review of ferrihydrite occurrence in soils (Childs, 1992) states that a small amount of

ferrihydrite (1-2%, generally below the level of detection by XRD) can have a large

effect on soil sorptive properties due to its high specific surface area and high

proportion of reactive sites. Ions such as Si sorbing on the exterior of iron oxides would

be readily available for exchange. The presence of impurity anions in the surrounding

solution can adsorb to the outer edges of ferrihydrite and inhibit crystallisation (Zhao et

al., 1994). Phosphate and organic anions may play a similar role in the inhibition of

crystallisation (Cornell, 1987) making arsenic a potential inhibitor based on its similar

adsorption affinity as phosphate.

The transformation of amorphous iron oxides to more crystalline phases can occur

rapidly (within hours), over several months or even years (Sorensen et al., 2000). Frau

and Ardau (2004) suggested the possibility of iron oxide aging contributing to arsenic

desorption in an environment where iron oxide tailings were discharged 35 years ago.

Aging may have an increased effect on arsenic sorbed to iron oxide coatings in the

“beach barrier sands” that were eroded prior to the “fluvial sands” (i.e, they may have

been mobilised into solution after aging). This could explain the lower correlation

between arsenic and iron in the “beach barrier sands” (r2 = 0.70) when compared to that

observed for the “fluvial sands” (r2 = 0.93) which are less re-worked and therefore not

as aged.

It thus remains that the distribution and presence of iron oxides in the matrix is difficult

to quantify, however, their presence has been identified and therefore their ability to

sequester arsenic is a process that may be further validated by geochemical modelling

and interpretation of aqueous data (Chapter 7).

6.3.2 Association with Pyrite

Minor pyrite was observed via XRD but major occurrences were not detected. A similar

case was observed by Fitzpatrick et al. (1996) where pyrite was not detected via XRD



but was later found under high powered microscope. SEM analysis of a “fluvial sand /

estuarine clay” sample located close to the estuary was conducted (ML7/18). Shell

material is abundant in this area, although this sample is predominantly a poorly sorted

fine-grained sandy olive clay containing both smooth and angular shaped pebbles. It

contains 10.4 mg kg-1. As in the sediment, while groundwater As was reported at 26 g

L-1 and 16 g L-1 over a one year interval between sampling events. The composition is

largely quartz (Si, O) with clay (Al, Mg, K, Fe, Ti) also potentially dominating. Various

surface aggregates were examined and generally found to be representative of the

remainder of the particle, with the exception of one analysis. A small spherical

aggregate (Figure 6.5) contained the elements noted above with the addition of sulfur

and an increase in iron. This implies that a sulfide mineral may be present sporadically

in the aquifer, however, no other SEM-EDX analyses reported sulfide.

Figure 6.5 Small agregate observed on a “fluvial sand / estuarine clay” sample. Clay and quartz are suspected however the presence of sulfur may indicate discrete sulfide mineral presence.

Further analysis of the potential for discrete sulfide mineral presence in the aquifer was

warranted. A separate “fluvial sand / estuarine clay” sample (ML7/12) was examined by

SX50 electron microprobe. Like sample ML7/18 this sediment is an olive grey sand

containing both lithic and shell fragments. Quantitative analysis of a shell fragment

produced atomic ratios indicative of calcite (Ca/O 0.32) with minor association of Fe

and Mg potentially via inclusion or sorption (Figure 6.6). A small bright cluster was

observed on a portion of the shell material. A quantitative analysis of this cluster

reported it was predominantly Fe and S, which suggested it was an iron sulfide,

potentially pyrite. The Fe/S ratio (which was previously 1.17 on the shell material) is



reported as 0.66 for this cluster, which is closer to the ratio for pyrite (0.5). The Fe

presumed to be substituting in the CaCO3 matrix is most likely contributing to this ratio,

thereby keeping it above the 0.5 ratio of pyrite. Arsenic is reported at 0.14%,

significantly above its DL of 0.02% and surprisingly high given its bulk XRF

concentration of only 5.8 mg kg-1. It is likely that arsenic has been included into the

pyrite crystal lattice by substitution of sulfur, which may also contribute to keeping the

Fe/S ratio above that indicative of pure pyrite. The presence of 0.14% Pb may indicate

cation association with pyrite by a sorption mechanism.

For comparison, electron microprobe analyses of arsenian pyrite in a gold mining

district of California produced an average As content between 0 and 0.09 wt.% for a

sample reporting a bulk concentration of 10 mg kg-1 (Savage et al., 2000). In addition,

up to 0.33 wt.% Pb was also reported in the arsenian pyrites of the region, with the trace

elements Zn, Mn, Sb, Se, Cu, Bi, Co, Ag and Au all generally reported below 0.1 wt.%.

Further, Thornburg and Sahai (2004) identified arsenian pyrite occurring as

disseminated grains, veins and nodules in sandstone of the Fox River Valley, Eastern

Wisconsin. Concentrations of arsenic in disseminated pyrite are comparable to those

reported herein – generally <1 wt%. The pyrite was initially formed due to migration of

ore forming Mississippi Valley type fluids.

To be consistent with the above cited literature, the presence of Pb in the Stuarts Point

arsenian pyrite may indicate formation in close proximity to a mineralised source in

comparison to authigenic precipitation in-situ. This will be further discussed during

analysis of pyrite formation in the arsenic mobilisation chapter.

Figu

re 6

.6 E

lect

ron

mic

ropr

obe

back

sca

tter

ed i

mag

e of

a s

hell

frag

men

t w

ith a

pyr

ite c

lust

er,

as d

eter

min

ed b

y th

e qu

antit

ativ

ean

alys

is.



Further sources of concentrated arsenic were found at this location. Figure 6.7 shows a

quartz grain (Si/O ratio 0.42) with some clay presence (Al, K, Mg and Ti). Arsenic and

Pb were both undetected.

The bright clusters observed on this grain were quantitatively probed for elemental

composition. While the Si concentration remained stable, O, Al, Mg and K all

decreased.

There is a significant increase in Fe and S – which was previously not detected. The

ratio of Fe/S in this cluster was calculated to be 1.25 which once again indicates an

abundance of Fe relative to S, however Fe has already been detected in the matrix of the

grain and is therefore assumed to be biasing this ratio.

In addition, arsenic was reported in higher concentration than the previous analysis

(0.37 wt.%) and may indicate increased substitution of arsenic for sulfur, which

contributes to the high Fe/S ratio. In an experiment showing precipitation of an arsenic-

iron-sulfur phase during bacterially mediated microcosm methods, the solid phase

exhibited a stoichiometry of 1:1:1 for Fe:As:S indicative of arsenopyrite formation

(Rittle et al., 1995). The ratio of Fe, S and O to As are all above 250, indicating the lack

of an arsenic bearing mineral such as arsenopyrite (FeAsS) or scorodite

(FeAsO4.2H2O).

Thus, discrete arsenian pyrite clusters are considered as the most likely arsenic-iron-

sulfide mineral phase present in the Stuarts Point aquifer matrix.

Figu

re 6

.7 B

ack

scat

tere

d el

ectr

on im

age

of a

qua

rtz

grai

n w

ith a

n ar

seni

c co

ntai

ning

pyr

ite c

lust

er o

ccur

ring

on

it.



6.3.3 Incorporation into clay minerals

A negative correlation observed between arsenic and clay-representative elements

suggests arsenic at Stuarts Point may not be associated with bedrock clay minerals.

Theoretically, however, it may be possible and should be investigated accordingly in the

event that the statistical correlations are not representative of true conditions. The

identification of illite was made via XRD and electron microprobe analysis. Through

both optical illumination and back scattered electron imaging, a coating was identified

on a selected grain in the matrix (Figure 6.8).

A

B

Figure 6.8 A coating on a grain in sample ML10/26.8 under electron microprobe (a) optical image (b) back scattered electron image. Scale identical.

A line scan was conducted to ascertain the compositional differences between coating

and grain. Eleven elements were chosen for the analysis: O, Mg, Al, Si, S, K, Ca, Ti, Fe,

Pb and As. Figure 6.9 shows the location of the line scan on the optical image and the

output graph to show changes in composition between coating and grain. Calculation of

atomic ratios indicate that the grain is composed of quartz (Si/O is predominantly 0.5)



and the layer coating the quartz grain is composed of illite clay (Si/O ratio is

predominantly within the range for illite as quoted by Velde [1992] – 0.26 to 0.31).

Minor fluctuations in composition occur within the non-homogenous clay coating,

including the presence of a quartz inclusion contained within the coating. No arsenic

was detected in either the illite coating or the quartz grain.

Figure 6.9 Line scan output chart and identified mineral compositions.



Illite has a moderate surface area available for sorption when compared to other clay

minerals such as kaolinite. Interpretation of aqueous conditions may aid in determining

arsenic association with illite, however analysis of the solid phase suggests it may

contribute insignificantly to arsenic retardation within the aquifer.

6.3.4 Sorption on Organic Matter

LOI can represent all components of the sediment which are easily oxidised. This

component may include carbonates, sulfides, water and organic matter content.

Therefore, using LOI as an indicator of organic matter content is not desirable. For

discussion purposes only, LOI results for the Stuarts Point sediments are low, ranging

from 0.3 to 6.2%, indicating that organic carbon percentage content is much less. The

lack of organic C can be attributed to sediments being deposited under barrier

conditions, fluvial sedimentation and a sub-tidal environment which did not support

mangrove habitats. For similarity, the Tweed River floodplain located north of the

Macleay River, contains low organic C (<20,000 mg kg-1) at 3-10m depth in fining

upwards delta and prodelta deposits (Lin et al., 1998). The organic matter deposited in

the Tweed River floodplain has been suggested to derive from either terrigenous

organic-rich soils or intertidal habitats (Lin et al.,1998). Similar conditions are proposed

for the Stuarts Point organic matter. Discontinuous layers of coffee rock, however, exist

at Stuarts Point and may contribute to arsenic sorption. Coffee rock was examined by

SEM and found to be composed of quartz with varying amounts of iron and aluminium

(Figure 6.10). Iron and aluminium may be indicative of grain coatings or cementation

layers spatially heterogenous within the coffee rock layer.

Figure 6.10 Coffee rock SEM photomicrograph.



The presence of Fe and Al oxide coatings may contribute to As immobilisation during

water table fluctuations. Direct correlation between arsenic and organic matter derived

from coffee rock is difficult to define due to the low number of coffee rock samples (n =

1). Arsenic concentrations in this sample are distinctively higher (8.3 mg kg-1) than the

average for the “beach barrier sands” (3.9 mg kg-1) however without further analyses

arsenic sorption within coffee rock layers can only be assumed.

6.3.5 Association with Calcite

Given the large amount of shell material encountered in the aquifer, specific attention

was placed on calcite as a potential sink or governing control on arsenic distribution.

Arsenic can substitute for carbonate in the calcite crystal structure. Small shell

fragments were analysed under SEM. These shell fragments ranged from approximately

500 m to large specimens several centimetres in size. Figure 6.11 shows a section of a

shell fragment at x100 magnification.

Figure 6.11 Portion of shell material identified as calcite with minor trace element impurities.

Ca/O ratios were indicative of calcite (0.34). Surface striations and small ‘growths’ are

clearly visible on the surface of the shell, indicating they are not merely homogenous

structures within the sediment matrix. A number of elements (Si, Al, K, Ba) were

reported as trace phases however arsenic was not detected.

Calcite and arsenic association was also examined with the electron microprobe. As

reported previously, arsenic was found associated only with arsenian pyrite which was



precipitated onto the calcite structure rather than isomorphically bound within the

calcite mineral.

6.3.6 Anion competition

Competitive adsorption between arsenic and phosphate has been suggested as a

probable process in the “bedrock clays”. A very strong correlation exists between

arsenic and phosphorous in this group and no other. Both the highest reported As and

P2O5 concentrations occur in the oxidised bedrock clay supporting potential anion

sorption onto iron oxides. The coffee rock sample contains similar ratios of arsenic and

P2O5 and therefore may also exhibit competitive sorption between the two anions. These

processes are further addressed in Chapter 7.

6.4 CHAPTER SUMMARY: IDENTIFICATION OF CONTROLLING SINKS IN THE AQUIFER

Theoretical and laboratory studies are useful in determining potential controlling sinks

present within the Stuarts Point aquifer based on their presence or absence, as discussed

above. Unfortunately, as Goh and Lim (2004) note, few studies have been conducted on

soils and sediments as a whole, which often contain a mixture of minerals and

geochemical reactions operating in disequilibrium. Each plausible arsenic sink has been

examined in relation to the Stuarts Point aquifer. Identification of some sinks have been

made, however direct evidence of arsenic association is not available for each one.

Given the characterisation of the aquifer matrix provided herein, the suggested arsenic

sources and sinks to be tested in the mobilisation chapter are (Figure 6.12):

Arsenic and phosphate are both sorbed onto iron oxide surfaces in the “bedrock

clays”. Illite is present but is considered a potentially minor arsenic sink, if at all.

Arsenic may be contributed to the depths of the aquifer via bedrock discharge

through mineralised zones or as a product of released arsenic after fluvial

transportation of gravels present within some clays at depth;

Fluvial and marine conditions exist intermittently hereafter, contributing arsenic

from the hinterland via deposition of arsenic sorbed onto iron oxides. Persistent



inundation by seawater after deposition produces an anaerobic environment which

dissolves these iron oxides, releasing arsenic which is re-precipitated as arsenian

pyrite under the reducing conditions. Alternatively, conditions may not dissolve the

iron oxides but simply reduce the adsorbed As(V) to As(III) which is then

preferentially released into solution for redistribution in a sulfide phase.

Alternatively, aging of iron oxides may promote desorption of arsenic for

redistribution or arsenian pyrite was transported fluvially rather than precipitated

authigencially leaving As in dissolved phase after its desorption;

The overlying “fluvial sands” were not subjected to persistent seawater inundation.

Therefore, arsenic sorption onto colloidal iron oxides or transported in dissolved

form via the Macleay River and subsequently adsorbed to iron oxides in the aquifer

matrix upon deposition, remains the dominant arsenic sink; and

Considerable re-working and homogeneity of the “beach barrier sands” leaves

arsenic sorbed to iron and/or aluminium oxides or associated with sporadic coffee

rock accumulation at the water table.

There are no abrupt boundaries between these arsenic geochemical controls, changes are

predicted to be gradational, contributing to the complex nature of arsenic

hydrogeochemistry. The association of arsenic with a number of sinks in different parts

of the aquifer causes its heterogenous distribution in dissolved form.

Figu

re 6

.12

Sum

mar

y of

As g

eoch

emic

al si

nks p

ropo

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here

in fo

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us S

tuar

ts P

oint

aqu

ifer.


7 ARSENIC MOBILISATION IN THE AQUIFER

7.1 GENERAL AQUIFER CONDITIONS

7.1.1 Major Hydrochemical Processes Groundwater is predominantly fresh (generally < 1,000 S cm-1) with slightly acidic-

neutral pH (4.7-7.9). Five dominant water groups exist in the Stuarts Point aquifer;

shallow groundwaters;

barrier sand groundwaters;

fluvial sand and clay groundwaters;

fresh/saline mixing zone; and

seawater intrusion groundwaters.

These groundwater groups were delineated by HCA and are plotted chemically on a

piper diagram, and vertically against a borelog, in Figure 7.1. The dendrogram is

located in Appendix B7. Shallow groundwater is characterised1 by Cl-HCO3-(SO4)-Na-

Ca type water and is representative of coastal wet/dry depositional processes with minor

oxidation of sulfides resulting in an average groundwater pH of 7.3 (Table 7.1).

Interaction with coffee rock at the zone of water table fluctuation may also contribute to

these lower pH values. The barrier sands are dominated by HCO3–(Cl)-Ca-Na

influenced by wet/dry depositional recharge waters filtering downwards through the

aquifer with shell material dissolving along flowpath. HCO3–Ca waters dominate most

of the aquifer in the fluvial sand and clay deposits. Towards the base of some bores a

mixing zone between fresh and saline waters is encountered characterised by a HCO3–

Cl-Na-Ca signature indicative of clay-water interaction and fresh/saline water mixing

before Cl-Na type waters dominate in the seawater intrusion zone. The Stuarts Point

groundwater composition is therefore largely controlled by both geochemical processes

(atmospheric input, sulfide oxidation, shell dissolution reactions, seawater intrusion)

and lithological controls (clay-water-shell interaction). Similar hydrochemistry has been

observed in coastal groundwaters of Western Netherlands (Stuyfzand, 1993).

1 Szczukariew-Priklonski hydrochemical classification (Alekin, 1970).

Figu

re 7

.1

Pipe

r di

agra

m s

how

ing

litho

logi

cal

and

geoc

hem

ical

in

fluen

ces

on m

ajor

ion

gro

undw

ater

com

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tion.

The

lo

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1 is

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ow g

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ally

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ofile

.

Chapter 7 – Arsenic Mobilisation in the Aquifer


Table 7.1 Mean concentrations of general parameters and major ions in each groundwater type identified at Stuarts Point (mg L-1 unless stated).

Groundwater Group

ECS/cm

EhMV

pH Na+ K+ Ca2+ Mg2+ HCO3- Cl- SO4

2-

Shallowgroundwaters

204 -25 7.3 17 5 10 4 19 32 22

Barrier Sand 387 -138 7.3 15 2 58 4 190 29 10 Fluvial Sand /

Clay 414 -109 7.3 15 3 59 5 167 30 27

Fresh saline mixing zone

1,240 -129 7.2 136 9 66 19 263 222 40

Seawater Intrusion

20,964 -200 7.2 3,632 102 404 461 283 6,953 890

Calcite dissolution (equation 7.1) is the dominant reaction responsible for the Ca-HCO3

nature of the groundwater. Saturation indices calculated in PHREEQCI (Version 2.11;

Parkhurst and Appelo, 1999) show that no groundwater samples (except one) are

oversaturated with respect to calcite (Figure 7.2) indicating calcite dissolution as a

likely process. The anomalous sample (ML10/5) is located directly below agricultural

land and therefore is suspected of being influenced by anthropogenic inputs.

CaCO3 + H2O + CO2 Ca2+ + 2HCO3- (7.1)

Figure 7.2 Saturation Indices for calcite with the various water groups separately graphed.



Dissolution of calcite can be expected due to the high content of shell material dispersed

throughout the aquifer and the observed disequilibrium between chemical composition

of the water with the solid phase. The shallow groundwaters are more undersaturated

due to their lower content of shell material. Figure 7.3 shows that all groundwater

samples plot on the 1:2 calcite dissolution line except the seawater intrusion samples

which have increased Ca2+ and Mg2+ concentrations due to mixing and ion exchange

reactions.

Figure 7.3 Ca+Mg+Sr versus HCO3- to indicate calcite dissolution. The excess

cations over HCO3- in the seawater intrusion samples is a result of

mixing and cation exchange processes.

Precipitation of gypsum is possible if calcite is dissolving concurrently with pyrite

oxidation; a phenomenon suggested for similar coastal sediments by Willett and Walker

(1982). The release of Ca2+ from calcite dissolution (equation 7.1) coupled with SO42-

release from pyrite oxidation can theoretically precipitate gypsum when concentrations

of Ca2+ and SO42- reach equilibrium (equation 7.2). Saturation indices calculated for

gypsum, however, indicate undersaturation and therefore gypsum precipitation is not

likely to be occurring at this time and is probably largely controlled by low SO42-

concentrations limiting its formation.



Ca2+ + SO42- CaSO4 (7.2)

Seawater intrusion is clearly an identifiable process (Figure 7.4). Samples statistically

defined as seawater intrusion and mixing zone samples fall on the seawater mixing line.

The increased Ca2+ and Mg2+ (noted previously) is at the expense of the exchanger Na+

(Figure 7.5a). Sodium is introduced to the system via seawater intrusion but is readily

exchanged on mineral surfaces for Ca2+ and Mg2+. Two moles of Na+ are exchanged for

one divalent cation of Ca2+ or Mg2+ in reverse ion exchange processes (equation 7.3).

The opposite is true for ion exchange processes (equation 7.4) which are occurring

during freshening of the aquifer (Figure 7.5b).

2Na+ + Ca-X 2Na-X + Ca2+ (reverse ion exchange – 7.3)

Ca2+ + 2Na-X Ca-X + 2Na+ (ion exchange – 7.4)

Where X is the mineral surface and Ca2+ is shown as the divalent exchanger.

Figure 7.4 Na+ versus Cl- shows seawater intrusion and mixing groundwaters plot on the seawater line (a). Other groundwaters (red box in {A} enlarged) are less influenced (b) by seawater composition.



Further information on regional groundwater chemistry can be found in Simpson

(2004), Smith et al. (2003) and Northey (2001). A full analysis of hydrochemical

controls in the Stuarts Point aquifer (outside the scope of this study) is concurrently

being prepared by the author herein for publication at a later date.

Figure 7.5 (a) Reverse ion exchange dominates the seawater intrusion samples whereas (b) the shallow groundwaters, barrier sands and fluvial sand and clay are subjected to both normal and reverse ion exchange processes.



7.2 ARSENIC GEOCHEMICAL PROCESSES

The contribution of arsenic by both lithological (e.g. arsenic sinks) and hydrochemical

(e.g. seawater intrusion) processes requires independent investigation. Therefore,

assessment of arsenic hydrogeochemistry will be conducted on six different

groundwater groups:

Shallow groundwater influenced by marine aerosol precipitation and water table

fluctuations (Group 1);

The high yielding sands of both barrier (Group 2) and fluvial (Group 3) origin,

differentiated based on transportation/weathering history;

The mixed fluvial sand / estuarine clay unit (Group 4);

Deep groundwaters from weathered bedrock clays (Group 5); and

Groundwaters affected by seawater intrusion (Group 6).

Groups 1 and 6 will deal with hydrochemical processes whereas groups 2-5 will assess

the plausibility of the arsenic sinks proposed in Chapter 6. Note that these groups differ

slightly from the HCA classified groundwater groups discussed above. The sands of

barrier and fluvial origin have been separated to account for arsenic variation due to

depositional differences; and the saline/fresh mixing zone has been included in the

seawater intrusion group to assess the overall arsenic relationship with seawater. The

average chemical composition for each of these ‘arsenic’ water groups is shown in

Table 7.2.

7.2.1 Arsenic Distribution and Speciation Dissolved arsenic concentrations range from 0.19 g L-1 to 83.69 g L-1 with mean

concentrations for each water group listed in Table 7.3. The previously high

concentrations (>330 g L-1) reported by Smith et al. (2003) indicate heterogeneity of

arsenic occurrence in the aquifer.

Table 7.2 Average chemical composition for each water group. Traces in g L-1

and majors in mg L-1.

shallow groundwaters

barrier sands

fluvial sands

fluvial sand/ estuarine

clay

bedrock clay

seawater intrusion

Ag 0.1 3.2 20.6 4.3 26.3 0.2Al 266.3 57.4 28.9 48.8 4.4 76.3As 6.4 4.8 10.9 10.7 43.3 39.5Ba 34.6 17.5 11.4 28.2 15.4 215.4Be 0.1 0.0 0.0 0.0 0.0 0.0B 0.0 17.0 19.5 122.4 49.5 0.0

Cd 0.0 1.6 23.5 11.7 0.0 3.9Ce 30.4 0.4 0.2 0.1 0.0 0.2Co 0.4 0.1 0.1 0.1 0.0 1.0Cr 2.0 2.0 1.5 1.4 1.2 3.8Cs 0.0 0.0 0.0 0.0 0.0 0.2Cu 0.0 0.2 0.0 1.3 0.0 2.8Dy 1.7 0.1 0.0 0.0 0.0 0.0Er 0.8 0.0 0.0 0.0 0.0 0.0Eu 0.7 0.0 0.0 0.0 0.0 0.0Ga 0.7 0.4 0.3 0.8 0.5 3.9Gd 3.1 0.1 0.0 0.0 0.0 0.0Ge 0.2 0.0 0.0 0.1 0.0 0.0Hg 0.0 0.2 0.3 4.7 0.0 0.0Ho 0.3 0.0 0.0 0.0 0.0 0.0La 21.3 0.2 0.1 0.1 0.0 0.1Li 1.8 1.1 1.0 5.8 2.1 89.4Lu 0.1 0.0 0.0 0.0 0.0 0.0Mn 24.1 84.4 46.7 60.9 28.8 1245.7Mo 0.1 1.6 0.3 0.9 0.7 1.5Nb 0.2 0.1 0.1 0.1 0.0 0.0Nd 20.2 0.3 0.1 0.1 0.0 0.1Ni 1.0 1.6 1.2 1.4 1.7 11.6P 30.5 83.3 112.4 139.6 44.9 172.1

Pb 3.2 5.1 4.3 3.0 10.4 1.9Pr 3.8 0.1 0.0 0.0 0.0 0.0Rb 10.1 4.0 3.1 3.5 2.2 32.2Sb 0.0 0.5 0.0 0.0 0.0 0.0Sc 1.4 1.2 1.6 2.1 2.6 3.5Se 0.5 0.4 0.4 2.0 0.4 42.4Sm 3.2 0.1 0.0 0.0 0.0 0.0Sn 0.0 0.3 0.2 1.7 0.0 0.1Sr 56.7 535.9 410.0 472.3 324.9 6901.4Ta 0.1 0.1 0.1 0.1 0.0 0.0Tb 0.3 0.0 0.0 0.0 0.0 0.0Th 0.1 0.0 0.0 0.0 0.0 0.0U 0.0 0.2 0.1 0.5 0.2 0.1V 1.4 1.9 1.2 1.1 0.0 23.6W 0.2 0.1 0.1 0.1 0.1 0.1Y 7.1 0.4 0.2 0.1 0.1 0.1

Yb 0.5 0.0 0.0 0.0 0.0 0.0Zn 30.3 2.1 4.1 2.0 1.2 5.4Zr 0.2 0.4 0.4 0.2 0.3 0.3

Sulfide 0.6 0.1 0.1 0.4 0.0 1.2Total Fe 1.6 1.3 1.2 0.7 0.0 11.8

Ca2+ 9.6 72.6 58.5 54.9 47.3 417.8Mg2+ 3.6 6.3 3.7 11.3 4.8 479.1Na+ 17.1 15.5 16.6 73.9 31.5 3752.9K+ 5.1 2.7 1.9 5.5 2.5 105.0

HCO3- 19.3 207.3 196.1 200.4 204.1 270.1

Cl- 31.7 31.9 29.5 120.8 33.2 7193.6SO4

2- 22.2 38.5 8.1 25.0 5.7 929.5PO4

3- 0.3 0.4 0.4 0.4 0.3 0.8EC 207.8 473.0 390.5 719.4 405.5 20042.2Eh -14.4 -122.7 -137.5 -111.7 -74.0 -243.3DO 1.0 0.5 0.6 0.7 0.7 0.1

Temp 20.9 20.9 21.2 21.6 21.2 20.5pH 5.3 7.3 7.4 7.3 7.5 7.1



Table 7.3 Descriptive statistics for dissolved arsenic concentrations in each lithologically or hydrochemically dominated water group ( g L-1).

Water Group n Minimum As Maximum As Mean As Shallow Groundwaters 19 0.2 30.8 6.5

Barrier Sands 47 0.5 20.3 4.8 Fluvial Sands 55 1.7 57.9 10.9

Fluvial Sand / Estuarine Clay 89 0.3 61.6 10.7 Seawater Intrusion 16 17.6 83.7 39.6

Bedrock Clay 2 30.6 57.9 43.3

Most groundwaters plot within the H3AsO30 stability range on a generalised Eh-pH

diagram for aqueous arsenic speciation (Figure 7.6). The few samples plotting within

the HAsO42- field represent the oxygen abundant estuarine surface water samples and

some shallow groundwaters.

Measurement of in-situ arsenic speciation was sought, however, the speciation methods

available had more disadvantages than advantages (Smith et al., 2003). The field site is

a considerable distance from analytical facilities raising concern over preservation of

arsenic species during transport to the laboratory. Bednar et al. (2002) explain a number

of factors that may affect arsenic species distribution:

Precipitation of metal oxyhydroxides caused by a change in redox or pH conditions

during sample collection – metal oxyhydroxides (particularly iron and manganese

oxyhydroxides in groundwater) provide sorption sites for dissolved arsenic species,

thereby altering species distribution after the sample has been removed from in-situ

conditions;

The oxidation of As(III) to As(V) – by photolytically produced free radicals

(particularly in iron containing water) can alter species distribution upon exposure

of the sample to solar radiation; and

The reduction of As(V) to As(III) – reduction can occur even after a sample has been

filtered to remove living organisms, also some forms of natural organic matter can

reduce As(V) to As(III).



Figure 7.6 Stuarts Point groundwater samples plotted on a generalised Eh-pH diagram for the As-O2-H2O system at 25ºC and 1 bar total pressure.

Despite some authors proposing suitable preservation methods (Bednar et al., 2002),

universally accepted preservation techniques for arsenic speciation were not in place

during the sampling period. An in-field aqueous speciation method was developed with

the Australian Nuclear Science and Technology Organisation (ANSTO) and trialled for

both these investigations and the speciation results reported in Smith et al. (2003).

The method used is the silver diethyldithiocarbamate method described by the

American Public Health Association (APHA, 1998). Custom-made field equipment was

designed to adapt the laboratory method to field conditions. Unfortunately, problems

with chemical instability under harsh Australian climates, combined with concurrent

laboratory studies assessing the reliability of the method (Crisp, pers. comm.) led the



author to reject the speciation results and raise concern over the values reported in

Smith et al. (2003).

Transport of a solute is largely controlled by its ability to be sorbed to aquifer materials.

Arsenite and arsenate exhibit variable sorption properties thus leading to separation of

dissolved As(III) and As(V) species during groundwater transport. Such

chromatographic separation along flowpath can lead to variable As(III)/As(V) ratios

throughout the aquifer and uncorrelate any original correlation between dissolved

phases (such as iron and arsenic) at the source (Smedley and Kinniburgh, 2002). This

makes interpretation of groundwater speciation data difficult. Arsenic speciation herein

will be assessed via solution parameters and geochemical modelling, bearing these

interpretative difficulties in mind.

7.2.2 Identification of Arsenic Mobilisation Processes Vertical distribution of dissolved arsenic in the multi-levels sampled for these

investigations are shown in Figure 7.7 and Figure 7.8.

Solid phase arsenic sinks were identified in Chapter 6. Table 7.4 lists the proposed

arsenic sinks corresponding to each water group. Low arsenic in dissolved phase may

indicate dominance by the proposed sink, i.e., arsenic remains in solid phase, explaining

the low concentrations observed in groundwater. However, arsenic geo-complexity may

further influence its distribution and hence each water group and corresponding arsenic

sink is examined in detail below.

Statistical methods were initially employed as a pre-screening tool to identify major

arsenic relationships in such a large data set. Pearsons correlations and HCA (Appendix

B7) identified the strongest arsenic-element relationships in each water group, which

were subsequently further assessed by hydrochemical interpretation methods. The

statistical pre-screening allowed the omission of large numbers of chemical bivariate

plots, which do not show evidence of arsenic geochemical relationships, thus keeping

this thesis concise and specific to the aims being addressed. Selected bivariate plots

exhibiting strong correlations are provided in Appendix B9 to show that these

relationships do exist, and that they are not the result of outliers biasing the statistical



result. It should also be noted that a lack of correlation may sometimes be the most

interesting result, such as the lack of correlation between iron and arsenic in the solid

phase of the fluvial sand/estuarine clays (Table 6.4). Calculation of saturation indices

(SI’s) and arsenic speciation was conducted in PHREEQCI. The results for each

separate water group are provided in Appendix B8.

Table 7.4 Proposed mobilisation processes (tested herein) from identified arsenic sinks at Stuarts Point and their influence on arsenic concentrations in groundwater.

Stuarts Point Water Group

Proposed Mobilisation Process from Solid Phase Arsenic Sink

Influence on Dissolved As

Concentrations Shallow

Groundwaters Oxidation of discrete arsenian pyrite and/or

coffee rock at the zone of water table fluctuation

Moderate localised effects

Barrier Sands Potential reductive dissolution or desorption from iron oxide coatings on sand grains

Extremely low concentrations

Fluvial Sands Reductive dissolution from iron oxide coatings on sand grains

Low to moderate concentrations

Fluvial Sand/ Estuarine Clay

Complete reductive dissolution of iron oxides; potential desorption from clays

High

Bedrock Clay Desorption from clay minerals and/or iron oxides

Moderate

SeawaterIntrusion

Suggested aqueous source of arsenic-rich estuarine water into the aquifer

Highest observed concentrations

7.2.2.1 Shallow Groundwaters The shallow groundwaters are characterised by a Cl-SO4-Na signature indicating

contribution by marine dry and wet deposition (Figure 7.9a). The low average pH of

the shallow groundwaters ( pH (shallow groundwaters) = 7.3) suggests iron sulfide oxidation

(equation 7.5) reactions may be contributing acidity, Fe2+ and SO42- in this zone of

groundwater fluctuation (Figure 7.9b, c and d).

FeS2 + 15/4O2 + 7/2H2O Fe(OH)3 + 2SO42- + 4H+ (7.5)

Figu

re 7

.7 V

ertic

al d

istr

ibut

ion

of d

isso

lved

ars

enic

for

ML

1 to

ML

5. S

tatis

tical

ly d

efin

ed w

ater

gro

ups a

re n

oted

on

each

gra

ph.

Figu

re 7

.8 V

ertic

al d

istr

ibut

ion

of d

isso

lved

ars

enic

for

ML

6 to

ML

10.

Stat

istic

ally

def

ined

wat

er g

roup

s are

not

ed o

n ea

ch g

raph

.



Figure 7.9 (a) Na+ versus Cl- (b) As Tot versus SO42- (c) Fe2+ versus SO4

2- and (d) AsTot versus Fe2+ concentrations in shallow groundwaters.

Sidle et al. (2001) found that the oxidation of arsenian pyrite in the Goose River

watershed of Maine was the source of arsenic in groundwater. However, oxidation

occurred under a slightly oxidising environment, which is in contrast to the Eh

conditions found at Stuarts Point. The slightly reducing waters and presence of sulfide

(approx. 1mg L-1) indicate arsenian pyrite oxidation by oxygen may not be plausible

under these geochemical conditions. Instead, arsenian pyrite oxidation by nitrate

reduction is proposed according to equation 7.6 and supported by the positive

correlations observed between Fe2+ and AsTot with NH4+ (Figure 7.10a and b).

8Fe(As,S)2 + 13NO3- + 25H2O + 10H+ 8Fe2+ + 8HAsO4

2- + 8SO42- + 13NH4

+ (7.6)



Figure 7.10 (a) arsenic versus ammonium in the shallow groundwaters, and (b) iron versus ammonium in the shallow groundwaters.

Nitrate-N is present up to 37.4 mg L-1 in the aquifer whereas DO averages <1 mg L-1

making the Eh values in the range suitable for nitrate reduction. Smith et al. (in press)

suggest this process does not drive arsenic enrichment in these groundwaters due to a

decrease in NH4+ with increasing arsenic concentrations observed in their bulk dataset.

However, thorough examination of shallow groundwater suggests this reaction is a

localised arsenic enrichment process controlled by anthropogenic NO3- presence in

shallow groundwaters in combination with discrete arsenian pyrite.

Ferric iron can also oxidise pyrite according to equation 7.7

FeS2 + 14Fe3+ + 8H2O 15Fe2+ + 2SO42- + 16H+ (7.7)

At pH values above 4 ferric iron is not available for pyrite oxidation due to the

insolubility of FeOOH (Appelo and Postma, 1999). Ferrihydrite, goethite and haematite

are reportedly saturated in samples ML2/5 and ML4/3; the only two samples to report

dissolved ferric iron concentrations. Formation of iron oxides and sorption of arsenic

are common by-products/processes of arsenian pyrite oxidation. The presence of iron

oxides is already pre-disposed in the aquifer during transport and subsequent deposition.

Some of these shallow groundwaters exhibit mildly oxidising conditions (up to +84

mV) and an overall correlation of r2= 0.7 between iron and arsenic in these shallow

sediments was observed suggesting potential adsorption of arsenic onto iron oxides in



localised areas. The coffee rock sample also reported more than 8% total Fe in the form

of iron oxides, which represents a localised ‘hot spot’ per se of iron oxide accumulation

in the shallow sediments. In contrast, calculation of arsenic speciation in PHREEQCI

indicates that As(III) as H3AsO30 is the dominant species, representing 100% of

modelled arsenic speciation. Adsorption of arsenite is greatest at pH values approaching

pH 9. The lower pH values observed for the shallow groundwaters may promote

conditions for arsenic occurrence in dissolved phase rather than adsorption to iron

oxides.

Oxidation of pyrite by nitrate reduction consumes acidity (equation 7.6). This acidity

may be supplied by the oxidation of pyrite by oxygen (equation 7.5) in localised zones

where slightly oxidising conditions have been observed. Alternatively, the low (sic) pH

observed for these groundwaters may be attributable to aluminium hydrolysis and

subsequent acid generation (equation 7.8)

Al3+(aq) + 3H2O Al(OH)3(s) + 3H+ (7.8)

Dissolved aluminium concentrations peak (Figure 7.11) in these shallow groundwaters

- Al(shallow groundwaters) = 266 g L-1; all other groundwaters have average concentrations

below 80 g L-1. Acid hydrolysis of the identified estuarine clay mineral illite may

provide dissolved aluminium for hydrolysis (Nriagu, 1978) (equation 7.9)

(K0.5Na0.36Ca0.05)(Al1.5Fe3+0.25Mg0.3)(Al0.45Si3.46)O10(OH)2 +7.41H+ + 2.59H2O

0.5K+ + 0.36Na+ + 0.05Ca2+ + 0.3Mg2+ + 0.25Fe(OH)3 + 1.95Al3+ + 3.46H4SiO4

(7.9)

however, clay content in the shallow groundwater zone is at its lowest (<1% as

represented by LOI). Therefore acid hydrolysis of illite in the shallow groundwaters is

unlikely to be a dominant process contributing dissolved aluminium to the water.



Figure 7.11 Plot of dissolved Al variation with depth for ML2. Statistical groundwater groups are noted also.

Instead, aluminium is suspected to be stored on exchange particles within the sediment,

notably coffee rock horizons. The influence of marine aerosol precipitation in this zone

of water table fluctuation provides Na+ ions which may be capable of exchanging with

aluminium sorbed on coffee rock constituents (equation 7.10)

coffee rock-Al(s) + 3Na+(aq) clay-Na(s) + Al3+ (7.10)

This dissolved aluminium is then available for hydrolysis and potential aluminium

oxide formation (equation 7.8). Saturation indices calculated in PHREEQCI show

saturation of boehemite (AlOOH), diaspore (AlOOH) and gibbsite (Al[OH]3) in these

shallow waters (Figure 7.12). All three minerals form from the weathering of

aluminous minerals (like illite) and are commonly found together in association with Fe

oxides (Nesse, 1991). Minor amounts of aluminium were reported in the aquifer

sediments with an observed correlation (r2 = 0.74) between arsenic and aluminium

reported for these shallow sediments. Diaspore was initially semi-qualitatively



identified by XRD; its presence in the aquifer can now be supported by the geochemical

modelling results. Arsenic can adsorb to aluminium oxide surfaces (Ladeira and

Ciminelli, 2004) and may be a controlling process in these shallow waters.

Figure 7.12 Saturation Indices of various aluminium oxide minerals (boehmite, diaspore and gibbsite) for the shallow groundwaters.

The localised coffee rock occurrences may contain humic or fulvic acids which are

capable of desorbing arsenic from iron oxides (Bauer and Blodau, 2005; Simeoni et al.,

2003) and potentially aluminium oxides; through competitive ion exchange on sorption

sites. This process may occur sporadically throughout the shallow groundwaters where

coffee rock horizons are present.

A strong correlation between AsTot, Cl-, EC, TDS and Li+ may indicate

evapoconcentration in shallow groundwaters of arid environments. Evaporation can be

dismissed in the temperate climate of Stuarts Point and therefore a correlation between

these elements is not expected and was not observed (Figure 7.13).

Thus the main process governing arsenic release in the shallow groundwaters is

suspected to be via arsenian pyrite oxidation by nitrate reduction. Arsenic may

subsequently be re-adsorbed onto aluminium and/or iron oxides thereby removing it



from dissolved phase. Given the low concentrations of As in both shallow sediments

and groundwaters, these processes do not appear to be dominant arsenic mobilisation

processes within the aquifer, but restricted to localised effects. However, NO3- addition

to groundwater (via fertilizer, soil zone processes or septic tanks, for example) where

arsenian pyrite is present may cause further As release and therefore should be

controlled.

Figure 7.13 AsTot versus (a) Cl- (b) EC (c) TDS and (d) Li+ in the shallow groundwaters.

7.2.2.2 Barrier Sand Groundwaters The mean sediment arsenic concentration for the “beach barrier sands” is 3.9 mg kg-1

and the mean concentration in aqueous solution is only 4.8 g L-1. Both values are the

lowest representative arsenic concentrations of any part of the aquifer and are typical of



average background values. Geochemical modelling suggests several arsenic

scavenging-minerals are saturated; pyrite, orpiment, goethite and various aluminium

oxides. Pyrite was observed in the underlying fluvial sand unit, however, oxidation of

pyrite is not supported by the aqueous data (Figure 7.14). If pyrite was oxidising, some

correlation between iron and sulfate would be visible (Figure 7.14a). If arsenic was

being released during the oxidation of pyrite, a correlation between arsenic and sulfate

could be observed (Figure 7.14b). Additionally, if nitrate was acting as the oxidant

instead of oxygen, a possible correlation between arsenic, iron and ammonium may be

observed (Figures 7.14 c and d).

Figure 7.14 (a) Iron versus arsenic (b) sulfate versus arsenic (c) arsenic versus ammonium and (d) iron versus ammonium in the barrier sand groundwaters.

The formation of orpiment (As2S3) can occur according to reaction 7.11



H3AsO30 + 1.5HS- + 1.5H+ = 0.5As2S3(s) + 3H2O (7.11)

Using thermodynamic data from Webster (1990) and substituting activity values

obtained from model solution species into equation 7.11, Ryu et al. (2002) obtained

positive orpiment SI values under very reducing conditions (<-170mV) in Owens Lake,

California. They concluded that, for the most concentrated arsenic sample reported in

their study (solid phase As 106 mg kg-1, dissolved As 96 mg L-1), orpiment may have

precipitated. Since the Stuarts Point aquifer contains significantly less arsenic and

sulfide concentrations in comparison to Ryu et al.’s (2002) study, orpiment precipitation

is not considered as a dominant arsenic sink for the barrier sands.

The presence of aluminium and iron oxides is suspected as coatings on quartz grains.

Adsorption onto these surface coatings may contribute to the small concentrations of

arsenic observed in the solid phase. This may be particularly true for barrier sands from

ML5 and ML6 which are the only groundwaters to exhibit higher As(V) concentrations

than As(III) concentrations (Appendix B8). There is no stratigraphic correlation for

these differences in speciation, however the high nitrate concentrations (16-33 mg L-1)

could be indicative of an anthropogenic agricultural source from the upgradient farm.

Consequently, high sulfate concentrations are also observed in this area. Dissolved

arsenic concentrations do not increase and may be controlled by the strong adsorption

affinity of As(V) for iron and aluminium oxides.

The trace concentrations of arsenic in the solid phase make verification of these arsenic-

mineral relationships difficult. XRD and electron microprobe techniques on

representative “beach barrier sand” samples did not identify any of these saturated

mineral phases. The above information suggests that the barrier sands do not provide

any distinctive dominant sink(s) or mobilisation process(es) for arsenic in the aquifer.

Slight arsenic peaks in these groundwaters may be derived from the downward

movement of shallow groundwater exposed to the oxidation of arsenian pyrite; or

dissolved arsenic derived from an upgradient reaction process having been transported

horizontally through the aquifer. Under these geochemical conditions, it is proposed that

the barrier sands represent good throughflow of groundwater with little potential for

retardation or further release of arsenic in groundwater.



7.2.2.3 Fluvial Sand Groundwaters An extremely strong correlation (r2 = 0.93) was observed between arsenic and total iron

in the “fluvial sand” sediments, thus leading the investigation towards the possibility of

arsenic sorption processes with iron oxides. Tareq et al. (2003) observed the opposite; a

very weak relationship between arsenic and total iron in Bangladeshi sediments was

attributed to the potential presence of insoluble Fe-silicates and minerals rather than Fe-

(hydr)oxides. However, they did report a good correlation (r2 = 0.68) between arsenic

and easily reducible iron in the sediments, indicating that arsenic may be associated

primarily with the amorphous iron oxide phase.

The excellent correlation between total iron and arsenic at Stuarts Point may indicate

dominance of iron oxide minerals rather than insoluble iron minerals. Oxides, rather

than any other iron-bearing minerals, were examined initially due to the orange-brown

colour of these sands which is often associated with their presence. Saturation Indices

calculated for the fluvial sand groundwaters indicate spatially variable saturation of

several iron oxide minerals; ferrihydrite, hematite, goethite, ferrosic hydroxide

(Fe3[OH]8)2 and lepidocrocite (FeOOH). Concentrations of dissolved ferric iron range

from 0 - 1.94 mg L-1 Fe3+ however ferrous iron concentrations are consistently higher

(up to 4.25 mg L-1 Fe2+) owing to the reducing nature of these waters (Eh (fluvial sand

groundwaters) range is –62 to –207 mV). Reductive dissolution of iron oxides and release of

bound arsenic has been proposed by several authors as the dominant mechanism of

arsenic release in Bangladeshi groundwaters (McArthur et al., 2004; Ravenscroft et al.,

2001; Nickson et al., 2000; Anawar et al., 2003; Tareq et al., 2003). The reaction is

catalysed by natural organic matter (equation 7.12),

8FeOOH + CH3COOH + 14H2CO3 8Fe2+ + 16HCO3- + 12H2O (7.12)

which promotes dissolution of the iron oxide and the subsequent release of any sorbed

arsenic. Differentiating between reductive dissolution and desorption from iron oxides

can be ambiguous under in-situ conditions. In a laboratory study by Guo et al. (1997)

arsenic was rapidly released after reductive dissolution of iron oxides suggesting it was

2 Langmuir (1997) reports that the infrequent identification of ferrosic hydroxide in nature may imply that the tabulated free energy of the phase is too negative.



dissolution rather than desorption governing arsenic mobility or that they occurred

simultaneously. Without such laboratory studies on the natural aquifer matrix, the

feasibility of iron oxide dissolution and/or desorption processes is best assessed by

examining solution pH, arsenic speciation and the relationship between Fe, HCO3- and

As concentrations in groundwater.

A good correlation between dissolved iron and bicarbonate in groundwater might be

expected if iron oxide dissolution is occurring (Figure 7.15a). Both ions, however, are

involved in other common geochemical reaction processes (e.g. silicate weathering, iron

sulfide precipitation) which may affect their concentrations in dissolved phase thereby

altering the expected relationship between the two ions. Iron oxides in the form of

ferrihydrite and goethite were both observed via SEM. A correlation between As and

HCO3- could also indicate arsenic release from iron oxide dissolution, however, the

addition of considerable amounts of HCO3- from calcite dissolution is one process

expected to mask this potential correlation (Figure 7.15b). Focussing instead on a

potential correlation between iron and arsenic (Figure 7.15c) does not provide any

additional clarity on the likelihood of iron oxide dissolution dominating arsenic

mobilisation. Both elements are highly susceptible to redistribution in separate mineral

phases thus altering any original correlation between the two elements. However, one

explanation is presented. Where total Fe concentrations are >0.03 mmol L-1 they are in

good correlation with HCO3- concentrations (Figure 7.15a) suggesting iron oxide

dissolution and potential arsenic release for these samples, yet Figure 7.15c indicates

arsenic concentrations are subsequently lower in these samples (i.e., As concentrations

are lowest when total iron concentrations are highest). This suggests that arsenic may be

re-adsorbed to freshly exposed iron oxide surfaces that have recently been dissolved, or

to other iron oxides in the aquifer which have not yet been dissolved. Arsenic

adsorption processes are considered to be rapid (Goh and Lim, 2004). The presence of

brown sediments in this unit supports the residence of iron oxides that have not yet been

dissolved; as does the optimum Eh range shown in Figure 7.15d. Arsenic

concentrations are highest from –50 to –150 mV before the onset of more reducing

conditions. These Eh conditions are more likely to be sulfate reducing waters and hence

conducive to arsenian pyrite formation, which will be further discussed shortly.



Figure 7.15 (a) Fe versus HCO3- (b) As versus HCO3

- (c) As versus Fe and (d) As versus Eh for the fluvial sand groundwaters.

A similar process of dissolution-readsorption has been proposed for Bangladeshi

groundwaters by McArthur et al. (2004) and may be responsible for maintaining low

arsenic concentrations in the dissolved phase of the fluvial sand groundwaters. The

driving force behind the reduction of iron oxides in Bangladesh is the oxidation of

organic matter. Peat layers distributed sporadically throughout the aquifers contain TOC

concentrations up to 34% (McArthur et al., 2004) which drives the reaction cited in

equation 7.12. Sand layers generally contain <1% TOC as reported in the same study.

For Stuarts Point, no peat lenses were encountered during the field investigations. Their

potential presence is inferred by reference to Hails (1968) who dated peat samples from



the inner sand barrier just south of Stuarts Point. Unfortunately, TOC concentrations

were not measured during these investigations due to time and cost constraints. LOI

provides an indication of organic matter content but also accounts for any oxidisable

matter present in the sediments, such as sulfides and dehydroxylation of clay minerals.

Therefore, the LOI concentrations measured for the Stuarts Point sediments (Table 7.5)

can only indicate that TOC is present no higher than LOI has been reported. Comparing

these results to those reported by McArthur et al. (2004) for Bangladesh, the Stuarts

Point sediments may contain minor amounts of organic matter to drive the reduction of

iron oxides. The heterogenous presence and nature of microbial populations throughout

the aquifer may also influence the composition of dissolved and particulate organic

matter (Schneider et al., 1994). Additionally, concentrations of DOM in most fresh

waters range from 1-20 mg L-1 and reach much higher values in wetlands (Abbt-Braun,

2002). The adjacent (upgradient) Yarrahapinni wetland could thus provide a source of

DOM to the aquifer.

Table 7.5 Mean and range for Stuarts Point LOI concentrations. Organic matter contents can be no higher than LOI.

Geomorphic Unit n % LOI range % LOI mean Beach Barrier sands 10 0.33 – 0.78 0.5

Fluvial sands 10 1.44 – 4.25 2.6 Fluvial sands / Estuarine Clays 13 1.27 – 6.19 3.2

Bedrock Clays 1 na 7.3

Thus far, the observed correlation between arsenic and iron in the “fluvial sand”

sediments suggests some arsenic remains associated with iron oxides in this part of the

aquifer. The geochemical data indicates reductive dissolution is occurring in the fluvial

sand groundwaters, but the identification of iron oxides in the solid phase shows

reductive dissolution has not yet gone to completion. Thus any arsenic which remains

bound to the iron oxide surface may also be available for:

redox transformations of bound arsenic;

desorption processes induced by pH or competing anions; and/or

aging of iron oxides promoting desorption.



These processes will now be assessed.

An additional reason for the poor correlation between iron and arsenic in Bangladesh

has been attributed to iron precipitation as siderite [FeCO3] and vivianite

[Fe3(PO4)2.8H2O] (Sracek et al., 2004; Bhattacharya et al., 2002). SI’s for these mineral

phases at Stuarts Point report that both phases are undersaturated in the fluvial sand

groundwaters, leaving iron in an aqueous phase. Neither were observed via SEM.

Furthermore, common Eh-pH diagrams for iron solids indicate that the siderite stability

field almost disappears when in equilibrium with goethite, suggesting the two minerals

rarely occur together (Langmuir, 1997). Since goethite has been identified in this part of

the aquifer, siderite is not expected.

As stated in the literature review, optimum As(III) adsorption onto both ferrihydrite and

goethite occurs at solution pH values between 6 and 9, the pH conditions encountered in

the fluvial sand groundwaters (Figure 7.16a). Additionally, the PZC for most HFO’s is

7.9-8.2 (Dzombak and Morel, 1990) with anion sorption promoted at pH values below

this; the pH of the fluvial sand groundwaters reaches a maximum of only 7.8, thus

adsorption of arsenic onto iron oxides is suited under these geochemical conditions.

The geochemical model calculated that 99.8% of arsenic was present as As(III)

however, the presence of As(V) may be expected due to thermodynamic disequilibrium.

The theoretical pH adsorption ranges and PZC values vary within the literature and as

such it is commonly accepted that As(III) is less strongly sorbed than As(V) under the

near-neutral pH of most waters given its uncharged state (H3AsO30). Accounting for this

decrease in electrostatic attraction and the assumed dominance of As(III), it is possible

that some arsenic is not strongly re-adsorbed after oxide dissolution thereby increasing

its presence in dissolved phase in this part of the aquifer.

Competition from other anions for sorption sites can also mobilise arsenic into

groundwater. Robertson (1994) found significant correlations (r2 = 0.56-0.87 at P<0.01)

between arsenic and Mo, Se, V, F and pH under closed basin conditions in the alluvial

basins of Arizona. He attributed these correlations to competition for adsorption sites

between these negatively charged anions in an aquifer where adsorption to iron oxides



was the dominant arsenic control. However, instead of replacing anions on adsorption

sites, the positive correlations observed by Robertson (1994) indicate that the chemical

similarity of these anions allowed them to be controlled (i.e., adsorbed) in a similar

manner. If adsorption sites become full, a negative correlation may be observed as

anions begin replacing other sorbed anions and releasing them into dissolved phase.

Phosphate is often examined due to its high negative charge and common occurrence in

groundwaters. Smith et al. (2002) found that PO43- presence significantly decreased

As(V) adsorption onto soils containing low amounts of Fe oxides, similar to the Stuarts

Point Fe oxide abundance. With As(V) concentrations assumed to be negligible in the

fluvial sand groundwaters, the effect of PO43- competition with As(III) is of more

concern. Goh and Lim (2004) found that PO43- had a more profound effect on sorption

capacity of both As(III) and As(V), when compared to the sorption competition from

other anions.

The poor correlation observed between arsenic and PO43- (Figure 7.16b) does not

support extensive competitive exchange between these two anions. Molybdenum and

arsenic concentrations exhibit a better correlation (Figure 7.16c) indicative of potential

for adsorption to the same sites, rather than competitive exchange processes. Indeed,

molybdate (MoO42-) has been found to inhibit arsenic adsorption onto mineral surfaces

(Manning and Goldberg, 1996). However, this relationship is expected to stem from the

surrounding molybdenite mineralisation, which has previously been shown to exhibit

arsenic as an associated element. Other common anions (Cr, V and U - Figure 7.16d-f)

show no apparent relationship with dissolved arsenic concentrations dismissing the

dominance of competitive interaction between arsenic and other anions.



Figure 7.16 (a) As versus pH (b) PO43- versus As (c) As versus Mo (d) As versus Cr

(e) As versus V (f) As versus U for the fluvial sand groundwaters.

Arsenic leaching by bicarbonate ions has been proposed for the Stuarts Point

groundwaters (Smith et al., 2006). It is difficult to delineate between an arsenic and



bicarbonate relationship derived from iron oxide dissolution versus desorption of

arsenic from iron oxide surfaces by bicarbonate ions; however it is indeed likely that

bicarbonate ions may contribute to arsenic mobilisation through surface complexation

mechanisms.

The dominant groundwater type for much of the aquifer is HCO3–Ca induced by the

dissolution of calcite. Bicarbonate concentrations average approximately 200 mg L-1 in

the aquifer. Szramek et al. (2004) proposed Sr could be an indicator of arsenic

occurrence in carbonate-rich groundwaters. In glacial drift aquifers of Michigan,

precipitation of carbonate minerals produced a correlation between Sr and As in older

groundwaters, indicating longer residence times.

No such correlation is observed for the Stuarts Point groundwaters (Figure 7.17),

potentially owing to the undersaturation of carbonate minerals. It is thus proposed that

the Stuarts Point groundwaters do not contain carbonate concentrations in such

abundance to enable the use of Sr as an indicator of As occurrence.

Figure 7.17 As versus Sr in the fluvial sand groundwaters of Stuarts Point.



Both aging of iron oxides and changes to the speciation of sorbed arsenic can promote

desorption. The aging of iron oxides leads to a loss in surface area when crystallisation

occurs (Frau and Ardau, 2004) and thus a loss in specific sorption sites. The

identification of both amorphous and crystalline iron oxides via SEM indicates iron

oxide crystallisation is a potential arsenic release mechanism in the aquifer. In addition,

with the onset of reducing conditions, any surface bound arsenic present as As(V) due

to redox disequilibrium, may be reduced to As(III) and subsequently mobilised.

Additionally, the formation of arsenian pyrite (solid phase arsenic and sulfur correlation

reported as r2 = 0.78 for the “fluvial sands”) may be occurring and will be discussed in

detail below. These processes are expected to contribute minor amounts of arsenic to the

fluvial sand groundwater and be overshadowed by the dominance of arsenic release

from reductive iron dissolution.

7.2.2.4 Fluvial Sand / Estuarine Clay Groundwaters The transition from “fluvial sand” to “fluvial sand / estuarine clay” is marked by a

change in sediment colour from orange/brown/red to olive green/grey. Often this change

in sediment colour is due to the disappearance of iron oxide minerals, which are

characteristically orange/brown/red. It can also be associated with a concurrent change

in redox potential from oxidising to reducing. The fluvial sand / estuarine clay

groundwaters show an increase in measured aqueous Eh ranging from –53 to –235 mV.

These geochemical conditions promote the complete dissolution of iron oxides, thereby

removing the potential for arsenic re-adsorption to fresh iron oxide surfaces. The

extremely strong correlation between iron and arsenic (r2 = 0.93) observed in the

overlying “fluvial sand” sediments is non-existent for the “fluvial sand / estuarine clay”

unit (r2 = 0.01). With the removal of iron oxide minerals for re-adsorption, arsenic

should remain in dissolved phase.

This process was observed by McArthur et al. (2004) for Bangladeshi groundwaters and

sediments. Where arsenic concentrations were >50 g L-1 reduction had gone to

completion and the sediments changed from brown to grey in colour. An increase in

dissolved arsenic in the fluvial sand / estuarine clay zone is noted for many of the

vertical profiles shown in Figure 7.7 and Figure 7.8. The average arsenic concentration



is 12.4 g L-1 but ranges from 0.3 – 61.9 g L-1. This is only a minor increase in

dissolved arsenic. However, sediment concentrations average 8 mg kg-1 which is also an

increase from 6.6 mg kg-1 As observed in the overlying “fluvial sand” unit. It appears

that an additional arsenic sink may be controlling arsenic distribution in this part of the

aquifer.

The presence of dissolved arsenic (0.30 – 61.9 g L-1), ferrous iron (0.14 – 3.88 mg L-

1), sulfide (0.0 – 2.85 mg L-1), reducing conditions (-53 to –235 mV) and the

identification of arsenian pyrite in this groundwater zone directs the interpretation

towards the likelihood of iron-(arsenic)-sulfide mineral formation and its resulting

stability under these geochemical conditions. The following questions are raised in this

discussion:

Are conditions currently suitable for arsenian pyrite formation in-situ? Alternatively,

Have conditions in the past been suitable for arsenian pyrite formation in-situ? Or,

Has the arsenian pyrite been transported from its original place of formation, i.e.

associated with mineralisation in the upper catchment?

Substantial evidence supporting either the first or second questions will confirm the role

of arsenian pyrite as a sink for arsenic, while confirmation of the third point would

consider it as a source of arsenic to the aquifer.

Geochemical modelling results indicate that pyrite is currently oversaturated in 93% of

the fluvial sand / estuarine clay groundwaters (Figure 7.18). With the presence of

dissolved arsenic already explained via the recent release from iron oxide dissolution,

arsenic may indeed be present in the groundwater for isomorphic substitution of sulfur

in the pyrite lattice. The formation of pyrite can be a complex process; Langmuir (1997)

notes that it does not simply precipitate directly from solution, rather it forms from the

successive sulfidation of a series of metastable Fe(II) sulfides. Schoonen and Barnes

(1991a; 1991b) list these sulfide sequences in more detail. Figure 7.19 shows

generalised pathways of pyrite formation, also described in equation 7.13

4SO42- + Fe2O3 + 8CH2O 2FeS2 + 8HCO3

- + 4H2O (7.13)



Figure 7.18 Saturation indices for pyrite in the fluvial sand / estuarine clay groundwaters. Ninety three percent of these waters are oversaturated.


Figure 7.19 Constituents required for pyrite formation (Appelo and Postma, 1999).

Aqueous redox potential is sufficiently reducing (up to –235 mV) to promote SO42-

reduction to S2- natural conditions. Marine estuarine clays may be the source of organic

matter and sulfate; with addition of iron and arsenic from the dissolution of iron oxides.

Once palaeowater salinity has exhausted the supply of sulfate pyrite precipitation may



be limited. As suggested in Figure 7.19 this process is bacterially mediated and is the

subject of much current research (Oremland and Stolz, 2005).

With the establishment that current geochemical conditions may be suitable for arsenian

pyrite formation, the possibility of past conditions forming pyrite (i.e., seawater

inundation of the aquifer during transgressions) do not need to be addressed in order to

define the likelihood of an in-situ formation of an arsenic sink. An alternate possibility

is the arsenian pyrite with its associated lead (and other potentially unknown trace

elements) could have formed in close proximity to a mineralised source in the upper

catchment and been transported by fluvial processes for later deposition. Detrital sulfide

minerals do not usually persist in the environment during erosion and transportation due

to oxidation processes. Craw and Chappell (1999) provide an example where such

detrital sulfide grains do persist – as a result of rapid erosion and transport providing

chemically immature sediments faster than oxidation can occur. In their investigation,

sediments are locally derived from arsenopyrite/pyrite metasediments within a few

hundred metres of their source; the Hyde-Macraes Shear Zone in New Zealand.

However, the length of transport via the Macleay River is in the order of kilometres, and

as such, preservation of discrete sulfide minerals over long distances is not likely at

Stuarts Point. Lead has also been identified associated with pyrite of acid sulfate soils

north of Stuarts Point (Smith and Melville, 2004), where source provenance (and thus

mineralisation) differs. In addition, the formation of arsenian pyrite on shell fragments

supports its designation as an in-situ arsenic sink, rather than source, for the Stuarts

Point aquifer.

Arsenic sorption and precipitation reactions on pyrite (in comparison to isomorphic

substitution) could also be occurring. Orpiment SI’s were calculated to be oversaturated

for the fluvial sand / estuarine clay groundwaters. The formation of an intermediate

iron-arsenic-sulfide precipitate on pyrite was observed by Bostick et al. (2004) and may

eventually lead to conversion of orpiment on the pyrite mineral surface. Molecular

studies of arsenic association with iron sulfide minerals is an emerging research area.

During the course of these investigations many different theories were produced,

including a recent study (O’Day et al., 2004) which provides spectroscopic evidences

that arsenic does not substitute for iron or sulfur in these minerals at the molecular scale.



The most that can be concluded at this stage is that arsenic is associated with iron

sulfide minerals at Stuarts Point, however, association via isomorphic substitution or

surface sorption/precipitation reactions cannot be differentiated. Detailed microscopic

investigations were not conducted as part of this study, but would provide beneficial

information on the association of arsenic with these mineral phases.

The highly reducing conditions are responsible for 99.9% of modelled arsenic species

being present as As(III). The speciation of arsenic (H3AsO30) is not conducive to strong

bonding to oxides and a lack of iron oxide minerals in this part of the aquifer decreases

the amount of re-adsorption of dissolved arsenic to any iron oxides remaining in the

aquifer matrix. Aluminium oxides may provide suitable sorption sites for some arsenic.

They exhibit an r2 value of 0.73 with arsenic in these sediments and modelling indicates

saturation of gibbsite and diaspore.

Additionally, the presence of illite may be conducive to arsenic adsorption. The

literature review noted the decrease in sorption affinity of As(III) compared to As(V)

but studies cited indicated illite (identified herein) was capable of both As(III) and (V)

sorption due to its high surface area. Maximum As(III) adsorption was found by

Manning and Goldberg (1997) to occur between pH 7.5 and 9.5 – the pH for these

groundwaters ranges from 6.6 to 7.8 and the highest dissolved arsenic concentrations

reported for these waters are at pH levels below this maximum adsorption range.

Therefore, some arsenic sorption onto illite may be occurring in the higher (sic) pH

waters. Less arsenic is expected to desorb over time due to diffusion of arsenic into the

lattice and strengthening of surface bonds.

The onset of estuarine conditions deposited many shell specimens in the sediments. The

association of arsenic with calcite was examined in Chapter 6 (sinks) finding no

apparent relationship via solid phase analytical techniques. The minor concentrations of

arsenic in solid phase did not identify any arsenic minerals via XRD; reported no

arsenic incorporated into calcite analysed by SEM and produced no appreciable arsenic

desorption via calcite digestion methods. Regardless, the sorption affinity of arsenic for

calcite has been documented in the literature. Given that all measured groundwater pH

values are below the PZC for calcite (approximately pH 10) it is plausible that arsenic



adsorption to calcite surfaces may be occurring. This process, however, is assumed to

have a negligible (if any) effect on dissolved arsenic concentrations throughout the

aquifer.

7.2.2.5 Bedrock Clay Groundwaters Only two sample points directly sampled groundwater from bedrock clay.

Concentrations of arsenic were highest (31 and 56 g L-1), however, no firm

conclusions can be made without a more representative data set. The decrease in Eh to –

74mV may indicate recharge from more oxidised bedrock groundwaters below.

Molybdenum derived from mineralisation has previously been discussed, and

additionally, concentrations of Ag (26 ug L-1) and Pb (10.4 g L-1) above natural

background concentrations and ecological guidelines3 confirms the theory that bedrock

groundwaters may contribute some arsenic to the unconsolidated aquifer. Other studies

have inferred arsenic contribution from bedrock; Warner (2001) found that arsenic and

chloride were positively correlated in groundwaters of the Lower Illinois River basin in

a manner similar to that observed between As and Mo for Stuarts Point. She concluded

that this correlation provided evidence of arsenic derivation from the underlying

bedrock.

The Pearsons correlations for arsenic in the solid phase bedrock clays were

predominantly negative with clay containing elements indicating a non-association with

arsenic and clay in the solid phase. However, only four sediment samples representing

these clays were analysed. This small sub-population may not be representative of all

bedrock clays and it is suspected that one sample, containing abundant iron oxides,

overshadowed the resulting statistics. Therefore it is proposed that groundwater with

arsenic derived from mineralised bedrock is discharging into the unconsolidated aquifer

where some dissolved arsenic may become adsorbed to illite or iron oxides depending

on their presence or absence at the discharge point. The modelled speciation indicated

dominance of As(III) therefore these processes are expected to be similar as those

described above.

3 ANZECC (2000) trigger values for freshwater (to ensure 95% species protection): Ag 0.05 g L-1 and Pb 3.4 g L-1.



7.2.2.6 Seawater Intrusion GroundwatersClearly, the seawater intrusion group contains elevated concentrations of dissolved

arsenic ( As(seawater intrusion groundwaters) = 39.5 g L-1). To demonstrate the correlation

between increased arsenic in the groundwater and the onset of seawater intrusion

processes, ML1 was chosen as a representative borehole (Figure 7.20). This multi-level

clearly differentiates each groundwater group discussed thus far when analysing

dissolved AsTot, Na+, Cl- and EC variation with depth.

Figure 7.20 AsTot, Na+, Cl- and EC variation with depth for ML1.

In brief review, arsenic increases slightly in the shallow groundwaters due to arsenian

pyrite oxidation. The influence of marine wet/dry deposition is also shown by a slight

increase in Na+, Cl- and EC. Arsenic concentrations in the barrier sand groundwaters

remain constant until reductive dissolution of iron oxides and release of sorbed arsenic



is responsible for the slight increase in arsenic concentrations in the fluvial sand

groundwaters. At the boundary between the oxidised “fluvial sands” and the reduced

“fluvial sand / estuarine clays” dissolved arsenic concentrations decrease suddenly due

to the formation of arsenian pyrite – a process sequestering arsenic in the solid phase of

this part of the aquifer. The subsequent gradual increase in Na+, Cl- and EC is

influenced by the presence of estuarine clays. At the seawater intrusion boundary these

constituents are at their greatest levels and co-incide with an increase in dissolved

arsenic concentrations.

Arsenic is extremely well correlated with both chloride and sulfate (Figure 7.21) in

these highly reducing groundwaters ( Eh(seawater intrusion groundwaters) = -243 mV). There are

two possibilities for this; geochemical conditions associated with seawater intrusion

promote enhanced mobilisation of arsenic from the solid phase, or the seawater itself is

a source of enriched arsenic concentrations.

Figure 7.21 Arsenic versus chloride (a) arsenic versus sulfate (b) in the seawater intrusion groundwaters.

The ionic strength of a solution can influence the degree of arsenic sorption due to

competition with background electrolytes for sorption sites. Outer-sphere metal

adsorption normally decreases with increasing ionic strength due to competition

(Bostick et al., 2003). Inner-sphere complexation is less affected by high ionic strength



and has been proposed as the dominant sorption mechanism between metal oxides

(Sherman and Randell, 2003) and many clay minerals (Manning and Goldberg, 1997).

Thus, the increased ionic strength due to seawater intrusion should not largely promote

As desorption in the aquifer. Additionally, iron oxide minerals are no longer dominant

in the reducing sediments and thus do not supply considerable sites for arsenic sorption

or desorption. This suggests that the change in geochemical conditions due to seawater

intrusion does not create an environment suited to increased arsenic release from the

aquifer matrix.

In fact, seawater intrusion conditions may actually promote some immobilisation of

arsenic in the aquifer. In a geochemical study of processes at the freshwater/seawater

interface in a shallow sandy aquifer in Denmark, Andersen et al. (2001) noted the

dominant redox process was sulfate reduction followed by methanogenesis. In the

seawater intrusion zone at Stuarts Point, the reduction of sulfate ( 2,196 mg L-1) to

sulfide ( 6.3 mg L-1) in the presence of dissolved Fe2+ ( 79.5 mg L-1) is conducive to

iron sulfide mineral formation and the possible sequestration of arsenic; a process

observed in the overlying fluvial sand / estuarine clay groundwaters. Strongly reducing

conditions and the oversaturation of iron sulfide minerals in the seawater intrusion

groundwaters supports these processes. The presence of ferrous iron, sulfide, arsenic

and organic matter in the absence of oxygen, is likely to increase the potential for

arsenian pyrite precipitation thereby removing arsenic from the dissolved phase. These

processes combine to indicate that seawater intrusion may actually immobilise dissolved

arsenic present in the aquifer. Therefore, seawater intrusion is not suspected to enhance

arsenic release into the groundwater. Instead, it is predicted to be an additional source

of arsenic to the Stuarts Point groundwater.

However, the decreased in dissolved arsenic due to arsenian pyrite formation observed

in the fluvial sand / estuarine clay groundwaters in Figure 7.20 is not apparent in the

seawater intrusion groundwaters. Instead, dissolved arsenic begins to increase with

increasing depth, suggesting an additional source of arsenic to the aquifer.

The adjacent Macleay River Estuary contains arsenic concentrations ranging from 23 –

79 g L-1 with a mean of 57 g L-1 (n = 15) which is significantly higher than the



groundwaters. Ashley et al. (2003) document the contribution of arsenic to the Macleay

River by oxidation of arsenic-rich stibnite deposits in the upper catchment (Figure 7.22)

while Tighe et al. (2004) note the contribution of arsenic via flood deposition (from the

same stibnite source) in the Macleay floodplain. It is proposed that dissolved arsenic is

being contributed to the aquifer via seawater intrusion of arsenic-enriched estuarine

water, however this requires further research4.


Figure 7.22 The Macleay River exhibits increased arsenic concentrations downgradient of the Hillgrove antimony mines (Ashley et al., 2003).

When this arsenic-rich seawater enters the aquifer, the highly reducing conditions

present in this part of the aquifer are sufficient to reduce As(V) estuarine surface waters

to As(III). All modelled arsenic is present as As(III) which further reduces the

likelihood of strong arsenic sorption by mineral surfaces. Likewise, the complexation of

4 Research is currently being conducted by the UNSW Groundwater Group to assess



arsenic in the seawater intrusion groundwaters is not expected to be a dominant reaction

since it is As(V) which is most strongly complexed with Ca2+ and Mg2+ in seawater

(Cullen and Reimer, 1989).

The combined effects of arsenic speciation, lack of iron oxide minerals and arsenic

behaviour in seawater are predicted to promote arsenic occurrence in dissolved phase.

Additionally, the formation of arsenian pyrite may be kinetically slow or have a limiting

component preventing further precipitation and arsenic sequestration. It is for these

reasons that arsenic in the seawater intrusion groundwaters is considered an additional

source of this toxic element to the aquifer.

7.3 CHAPTER SUMMARY

Dissolved arsenic varies in the Stuarts Point coastal aquifer due to a combination of

both lithological and geochemical processes. The heterogeneity of the aquifer sediments

causes redox stratification to occur, which in turn governs arsenic mobility in the

groundwater. Desorption, re-sorption and precipitation of arsenic occurs within the bulk

of the aquifer and is a consequence of changes in redox conditions. Bose and Sharma

(2002) proposed that,

“substantial (arsenic) mobilization is thought to occur under shifting redox

conditions” (pg 4917)

which appears applicable to the Stuarts Point groundwaters as it experiences changes to

redox conditions with increasing depth.

Natural and anthropogenic effects contribute to arsenic mobilisation processes. Natural

effects include water table fluctuation inducing geochemical changes in the shallow

system and aquifer heterogeneity contributing to changes in aqueous equilibrium.

Anthropogenic nitrate addition at the ground surface increases arsenic mobilisation in

shallow groundwaters. Seawater intrusion contributes further dissolved arsenic to the

groundwater system from arsenic-enriched estuarine water potentially derived from the

upstream erosion of arsenic-bearing mineral deposits and related mine tailings.

arsenic distribution in the Macleay Estuary and surface-groundwater processes.



The previous mobilisation processes proposed by Smith et al. (2003) have been

superceded by the more detailed analyses contained herein. The two principle processes

listed by Smith et al. (i.e., dissolution of Al hydroxides and release of adsorbed As(V)

or concurrently via pH-influenced desorption of As enriched Fe oxides; and the leaching

of the aquifer’s sandy-clayey matrix by groundwaters rich in HCO3- and of high pH) are

suggested as minor arsenic control mechanisms. Instead, the reductive dissolution of

iron oxides and precipitation of iron sulfide minerals has been shown to dominate

arsenic mobilisation at Stuarts Point.


8 CONCLUSIONS AND RECOMMENDATIONS

8.1 THE PROBLEM

Elevated arsenic has been identified at Stuarts on the mid-north coast of NSW. Stuarts

Point is a small coastal town located on the floodplain of the Macleay River and home

to several hundred local residents and farmers. Concern was raised over the presence of

such high concentrations in a drinking water supply aquifer. This study investigated the

source, geochemical sinks and aqueous mobilisation processes governing arsenic

distribution in the aquifer. Arsenic occurrence in a sandy aquifer was considered

unusual, since no source of geogenic arsenic was obvious, and few aquifers of this

nature had previously been investigated for arsenic. The findings of this research are

thus unique, and are presented below.

8.2 PROPOSED ARSENIC GEOCHEMICAL MODEL FOR THE STUARTS POINT AQUIFER

Figure 8.1 shows the major processes and inter-relationships, as proposed herein, for

the occurrence and mobilisation of arsenic within the Stuarts Point aquifer. This model

is described below.

Mineralised bedrock groundwaters contribute arsenic to the overlying weathered

bedrock clays.

The wet climate of the Pleistocene and Holocene promotes weathering in the

upper Macleay River catchment.

Natural stibnite deposits containing arsenic and antimony are weathered to form

iron oxides.

These oxides are present in colloidal forms and as coatings on sediment grains,

which are transported downstream and deposited to form the aquifer matrix.

Chapter 8 – Conclusions and Recommendations


Groundwater residence time, depositional history and age of the sediments

enables redox stratification to occur in the aquifer.

The deeper, more reducing fluvial sand and estuarine clay groundwaters have

undergone complete reductive dissolution of iron oxides resulting in the

subsequent mobilisation of arsenic into groundwater.

Some of this arsenic has been incorporated into iron sulfide mineral precipitates,

forming current arsenian pyrite sinks within the aquifer.

The fluvial sand sediments above have not completed reduction of all iron

oxides within the sediments, causing re-adsorption of arsenic to the remaining

iron oxide surfaces. This helps to maintain low arsenic concentrations in the

groundwater.

Conditions are not suitably reducing enough in the overlying barrier sands to

promote abundant iron oxide dissolution, which is responsible for the

background arsenic concentrations observed in this part of the aquifer.

Shallow groundwaters are exposed to nitrate input from the ground surface and

the presence of organic matter in the form of coffee rock lenses. Nitrate

contributes to the oxidative release of discrete arsenian pyrite phases while

DOM potentially influences sorption of arsenic to oxide surfaces.

The current extraction of groundwater from the aquifer induces seawater

intrusion of arsenic-rich estuarine water, bringing further dissolved arsenic into

the aquifer.

Figu

re 8

.1 A

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ic so

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and

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mod

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.



8.3 CONCLUSIONS

The main conclusions, addressing the proposed hypotheses from Chapter 1, are listed in

below:

1. Where is the arsenic coming from ?

Arsenic is derived from regional erosion of As-rich stibnite deposits that

have been fluvially transported and deposited to form sediments of the

Stuarts Point aquifer. Some minor arsenic is locally contributed from

mineralised bedrock sources.

2. What form is it taking in the aquifer matrix ?

Arsenic sorbed to iron oxides during fluvial transport now form part of

the aquifer matrix. Disseminated arsenian pyrite grains have been

identified within the aquifer matrix and are suspected of forming after

deposition of the aquifer. During their formation they sequestered

dissolved arsenic into their crystal structure. The presence of minor

aluminium oxide, humic matter and illite clays are suspected of retarding

arsenic under suitable, localised, geochemical conditions.

3. How is it being released into groundwater ?

Geochemical conditions, dominated by changes in the redox state of

groundwaters, produces several arsenic release and retardation

mechanisms in the aquifer. Processes include the oxidation of arsenian

pyrite in the shallow system; reductive dissolution of iron oxides;

desorption from mineral surfaces; and the introduction of increased

arsenic concentrations via arsenic-enriched seawater intrusion.

These findings make a solid contribution to the science and management of arsenic

geochemistry in the following manner:



Previously proposed arsenic sources for this aquifer differ from those discussed

herein. The source of arsenic has been successfully deduced and supported with

sufficient evidence. The derivation of a natural arsenic source reduces any potential

blame for anthropogenic introduction concerning its initial presence in this aquifer.

Prior to these investigations, the geochemical processes responsible for arsenic

distribution in the Stuarts Point aquifer (both solid and aqueous forms) could only

be postulated from the limited aqueous data available. Complex arsenic geochemical

processes have been delineated in great detail herein, and confirmed and/or

dismissed with a more robust geochemical data set.

A greater understanding of the source, retardation and mobilisation of arsenic in this

aquifer contributes to our broad understanding of arsenic in the environment; and

allows aquifer specific management procedures and research recommendations to be

made.

Previous studies suggesting arsenic in the Stuarts Point aquifer is derived from

adsorption to clays during sea level rise have been questioned. The contribution of

arsenic onto oxide and clay surfaces from seawater during marine transgressions has

been proposed for Bangladesh, however the absence of an arsenic-rich volcanic ash

deposit over which the Bangladeshi sea transgressed reduces the plausibility of this

theory occurring at Stuarts Point.

The suggestion that groundwater arsenic occurrence has implications for the

management of Australian coastal aquifers remains, but not solely due to the presence

of ASS or exposure of these coastal sediments to Quaternary sea level fluctuations.

Rather, any coastal or unconsolidated aquifer that has sediments derived from

mineralised provenances should consider monitoring for arsenic, and other potentially

toxic trace elements, in their groundwater systems.



8.4 RECOMMENDATIONS

8.4.1 Aquifer Specific (Stuarts Point)

A number of unanswered points have been raised during the course of this research.

Some of these are listed here to provide guidance on possible mitigating measures for

reducing the arsenic impact at Stuarts Point. They should not be relied upon as a

complete solution to the arsenic problem and should only be implemented if required,

after further consideration and investigation.

Management of aquifer redox equilibria – minimising anthropogenic additions

Oxidation of arsenian pyrite by nitrate reduction has been identified as a potential

arsenic mobilisation process in the aquifer. Arsenian pyrite may be oxidised by other

reductants (such as oxygen), however, nitrate addition to groundwater, particularly

shallow waters, should be minimised to reduce arsenic release. Examination of nitrate

inputs to the aquifer should include contributions from fertilizers (e.g. manure),

degrading (landfill) waste and septic tanks.

The addition of organic matter may also enhance the reductive dissolution of iron

oxides and the subsequent release of arsenic into the groundwater. Anthropogenic

organic matter input may include human and animal organic wastes; mulching or

fertilizer input; and leachate from landfills. Changes in redox equilibria brought on by

such examples should be minimised.

Examining the redox driver

Analysis of Total Organic Carbon (TOC) in the Stuarts Point aquifer sediments would

provide a better indication of processes driving the reduction of iron oxides and release

of arsenic into groundwater. Likewise, the identification of peat lenses in the aquifer

may provide information on areas where enhanced iron reduction and arsenic release

may occur. Further in-situ studies on the mobilisation of arsenic in this aquifer would

benefit other aquifers experiencing similar geochemical conditions, and provide further

management options for the aquifer. Experiments such as the injection of molasses,

nitrate and low arsenic water by Harvey et al. (2002) would allow more detailed

examination of the redox driver within the Stuarts Point aquifer.



Impacts on the environment

Although not reported herein, crops irrigated with untreated groundwater may be at risk

of trace element accumulation. Regular testing of vegetation should be maintained. The

detection of dissolved arsenic at elevated concentrations in the estuary requires further

research; delineation of the actual source (potentially from an upstream mining area)

should be identified from both a mitigation and possibly a legal viewpoint. In the

interim, measures should be in place to ensure protection of the environment,

particularly marine organisms which may bioaccumulate arsenic. Consumption of

organisms from the estuary by humans should be avoided unless given the all clear.

Seawater intrusion into the freshwater aquifer should be effectively managed to

minimise further arsenic input to the groundwater.

Groundwater extraction

The identification of discrete arsenian pyrite phases within the aquifer matrix warrants

further investigation into the effects of oxidation on arsenic release into groundwater.

Thornburg and Sahai (2004) noted that arsenic concentrations in groundwater affected

by discrete arsenian pyrite phases peaked after one hour of pumping or exposure to air

during laboratory investigations. Pyrite was presumed to be oxidised during this time,

releasing arsenic prior to the formation of surface iron oxide precipitates which

subsequently re-sorbed arsenic species, thereby lowering concentrations in dissolved

phase. These findings may have implications for groundwater extraction from the

Stuarts Point aquifer. Arsenic concentrations during well pumping should be monitored

to assess changes in dissolved arsenic concentrations. If arsenic decreases over a

specified pumping time, groundwater extracted for domestic and irrigation use may be

of a more suitable quality if it is pumped before consumption / application.

Alternatively, this may act as a relatively inexpensive pre-treatment mechanism leading

to a reduction in chemical treatment currently being utilised by local council. Therefore,

a simple pump test measuring arsenic concentrations during pumping could provide

improved management options for the aquifer.

Temporal uncertainties

Investigations herein and those reported in Smith et al. (2003) indicate that arsenic is

highly variable in the groundwater at Stuarts Point. Sidle (2002) found that arsenic



concentrations in the Goose River basin in Maine derived from the oxidation of arsenian

pyrite were highly variable over a two year sampling period. No specific temporal

investigations have been carried out at Stuarts Point and as such it should be recognised

that arsenic concentrations in groundwater may be prone to variation in both spatial

distribution and observed concentration ranges. For a better understanding of such

processes temporal studies would be beneficial.

Formation of discrete mineral phases capable of retarding arsenic mobility

Geochemical modelling outputs often produced results indicating conditions were

suitable for formation of arsenic minerals, or minerals capable of capturing arsenic from

solution. The presence of trace amounts of arsenic in the solid phase makes verification

of these potential minerals difficult. More detailed studies of the solid phase of the

aquifer matrix may enable discrete micron-sized arsenic bearing and/or ‘capturing’

minerals to be confirmed or dismissed. The potential presence of such small mineral

phases in a predominantly sandy medium may have implications for trace element

mobility in high yielding unconsolidated aquifers.

Analysis of antimony occurrence in the aquifer matrix

Antimony geochemistry, in general, is lacking in the literature. The few articles that are

available on antimony occurrence and chemistry suggest it is present in a solid

immobile phase. The elevated levels of antimony present in the Stuarts Point aquifer

matrix were thus attributed to an unknown solid antimony phase. Research on the

aquifer matrix, perhaps high resolution microscopy such as EXAFS, would confirm or

deny the presence of such antimony phase. In addition, further research on antimony

geochemistry would greatly benefit the small amount of research that is currently

available.

Further geomorphic work

Dating of sediments and shells may provide more information on the age of sediments

at their deposition. This would help to better define the geomorphic history of the

aquifer and extrapolate these findings to sea level changes elsewhere.



8.4.2 Other Research Recommendations

Advances in the science of arsenic geochemistry can be beneficial in predicting its

mobilisation/retardation at Stuarts Point. The following list provides examples of

technical research that may further aid in the interpretation of arsenic geochemical

processes in this, and other, aquifers. Some general conclusions drawn herein (not

necessarily arsenic-related) are also mentioned as their findings may be applicable to

other aquifers.

Advancement in the micron scale investigation of arsenic partitioning in solid solution

Much of the current arsenic research deals with its behaviour at the mineral surface.

Further increased knowledge of arsenic partitioning in various mineral phases (i.e.,

specific bonding mechanisms); mineral stability; and the effects of other system

components (ionic strength, competitive anions, speciation, aging) can improve

predictions on its likely mobilisation and bioavailability in the environment.

Application of theoretical knowledge to natural aquifer sediments

The investigations noted above are generally conducted on synthetic materials under

controlled laboratory conditions. This is essential to our understanding of the underlying

mechanisms but is not representative of true aquifer conditions. Different studies yield

different results, as can synthetic materials versus natural ones. For example, aquifer

sediments contain numerous minerals all potentially contributing to the sorption of

arsenic. The combined analysis of surface titrations to determine bulk PZC of an

aquifer’s clay and sand fractions, with batch experiments to characterise how each

fraction sorbs arsenic, would allow better judgement to be made regarding arsenic

sorption properties (i.e., outer-sphere versus inner-sphere complexation). In turn, this

would allow more confident predictions to be made about the dominant mechanism of

arsenic sorption in the aquifer and its likely potential for mobilisation. As such,

concurrent laboratory investigations with actual aquifer sediments would be particularly

beneficial prior to geochemical modelling or remediation trials. Research such as this

would help to improve the overall discipline of geochemistry.

Potential for arsenic occurrence in other coastal aquifers of NSW



The original report identifying elevated arsenic in the Stuarts Point aquifer (Piscopo,

1996) also identified dissolved arsenic concentrations in other coastal aquifers of NSW.

Stuarts Point was given priority for further investigation due to use of its groundwater

for human consumption. Deduction of a hinterland derived arsenic source for the Stuarts

Point aquifer could have implications for other unconsolidated aquifers deposited by the

Macleay River fluvial depositional environment. Additionally, other large river systems

present in NSW, such as the Manning River, may have had the potential to erode

hinterland arsenic sources and thus deposit aquifers conducive to arsenic mobilisation.

Alternatively, pyrite formation in the presence of arsenic (whether it be a natural or

anthropogenic source) followed by oxidation can lead to elevated arsenic in both coastal

and inland groundwaters. Investigation of other fluvially deposited aquifer sediments of

the Macleay River is recommended, followed by examination of fluvially deposited

coastal sediments derived from other river systems, especially if a hinterland arsenic

source is suspected. Additionally, the potential for arsenic incorporation into sulfide

minerals followed by oxidation, should be incorporated into risk assessment procedures.

Association of trace elements with pyrite and acid sulfate soils

The association of trace elements with pyrite was documented in the literature review.

In the Goose River Basin of Maine, a groundwater environment where arsenic is

proposed to originate from the oxidation of arsenian pyrite, analysis of pyrites in the

area reported up to 1.93% As; 0.02% Cu; 0.08% Pb; 0.01% Bi; 0.06% Ag; 0.34%

Co; 0.09% Ni; 0.03% Au and 0.02% Sb (Sidle et al., 2001). Unfortunately, their

relative concentrations in groundwater were not reported.

The inclusion of arsenic into the pyrite structure is well documented, however the

association of other trace elements has been given much less attention. Investigations

into trace element associations with pyrite formed under different environments may aid

in predicting the risks associated with pyrite oxidation. For example, pyrite at Stuarts

Point was found to contain both arsenic and lead, possibly as a result of their presence

in the sediments derived from a mineralised provenance. However, lead and copper has

been shown to be scavenged by iron monosulfide formation in sediments of the Tweed



River floodplain (Smith and Melville, 2004) indicating that provenance may not play an

important role in the association of trace elements with iron sulfide minerals.

Additionally, saline sulfidic soils developed under dryland salinity conditions inland

(Fitzpatrick et al., 1996) may also contain trace elements incorporated in their crystal

structure. The association of trace elements with pyrite should be the focus of additional

research, particularly in relation to their potential environmental impact and health risk

upon oxidation of ASS.

Mechanisms of pyrite oxidation by nitrate reduction

While pyrite oxidation by nitrate reduction is often cited as a potential arsenic

mobilisation mechanism, few reports of its occurrence in the environment have been

made. Data for the shallow groundwaters of the Stuarts Point aquifer indicate this

reaction may be occurring. Inputs of nitrate from anthropogenic activities such as

farming, may lead to pyrite oxidation in shallow groundwaters. This may be of

increased significance to acid sulfate soil environments where land is often used for

farming or settlement, potentially increasing the likelihood of nitrate addition to the

subsurface.

Heterogeneity in coastal groundwaters – source provenance and redox stratification

Groundwater drawn from ‘clean’ aquifer sands should not automatically be assumed as

‘safe’ for drinking and irrigation water supply. Although sparsely populated, with no

major source of contamination and producing water from ‘clean’ aquifer sands; the

Stuarts Point groundwater contains arsenic, and some other elevated elements, as a

product of its source provenance. Heterogeneity in aquifer conditions (such as changes

in redox) can cause these elements to be elevated in different areas of the aquifer.

Groundwater in similar environments being used for domestic or irrigation water supply

should be thoroughly investigated, addressing chemical and sediment heterogeneity,

before its use can be regarded as safe. The importance of detailed vertical studies to

show aquifer heterogeneity is particularly emphasised.

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Zheng, Y., Stute, M. van Geen, A., Gavrieli, I., Dhar, R., Simpson, H.J., Schlosser, P. and Ahmed, K.M. 2004. Redox control of arsenic mobilization in Bangladesh groundwater. Applied Geochemistry, 19, 201-214.

Zhu, Y., Zhang, X., Xil, Q., Chen, Y., Wang, D., Liang, Y. and Lu, J. 2005. Solubility and stability of barium arsenate and barium hydrogen arsenate at 25ºC. Journal of Hazardous Materials, 120, 37-44.


References

Zobell, C.E. 1946. Studies of the redox potential of marine sediments. Bulletin of the American Association of Petroleum Geologists (AAPG), 30 (4), 477-513.


APPENDIX A

Publications

O’Shea (2006) Page A-1

PUBLICATIONS RESULTING FROM THIS THESIS

Conference Abstracts and Papers (* peer reviewed) O’Shea, B., Smith, J.V.S. and Jankowski, J. 2001. The correlation between arsenic,

pyrite and marine clays in a coastal aquifer. Arsenic in the Asia-Pacific Region

Workshop, Nov 20-23, Adelaide, Australia, 36-38.*

O’Shea, B. and Jankowski, J. 2002a. Irrigation conflicts in a coastal aquifer – risks from

natural arsenic cycling. Irrigation Australia, May 20-23, Sydney, Australia.

O’Shea, B. and Jankowski, J. 2002b. Natural arsenic in a coastal aquifer: implications

for water quality management in Australian coastal zones. 27th Hydrology and Water

Resources Symposium, May 20-23, Melbourne, Australia, 117.*

O’Shea, B. and Jankowski, J. 2002c. Anthropogenic and geomorphic factors

contributing to arsenic distribution in a coastal aquifer, New South Wales, Australia.

XXXII IAH & VI ALHSUD Congress, Groundwater and Human Development,

October 21-25, Mar del Plata, Argentina, 110-116.*

O’Shea, B. and Jankowski, J. 2003. The use of solid phase selective extraction

techniques to support groundwater chemical data from a coastal aquifer affected by

elevated arsenic concentrations. 7th International Conference on the Biogeochemistry

of Trace Elements, June 15-19, Uppsala, Sweden, Volume 1, Scientific Programs II,

43-44.*

O’Shea, B. and Jankowski, J. 2004. Uncovering the source and distribution of elevated

arsenic in a coastal aquifer: Clues from the aquifer matrix. Inaugural Hydrogeology

Conference, December 1-3, Melbourne, Australia.

O’Shea, B. and Jankowski, J. 2005. Arsenic cycling in a Coastal Aquifer: Uncovering

the Source, Sinks and Mobilisation Processes. 8th International Conference on the

Biogeochemistry of Trace Elements, April 3-7, Adelaide, Australia.*

O’Shea, B. and Jankowski, J. 2005. Arsenic distribution in a coastal aquifer affected by

seawater intrusion. Geological Society of America Annual Meeting, 16-19 October,

Salt Lake City, Utah, USA. Abstracts with Programs, 37, (7), 170. *

O’Shea, B. and Jankowski, J. 2006. Naturally elevated arsenic and other trace elements

in a sand aquifer of eastern Australia. 41st Annual Meeting of the Northeastern

Section, The Geological Society of America, March 20-22, Camp Hill/Harrisburg,

Pennsylvania, USA. Abstracts with Programs, 38, (2), 88. *

O’Shea (2006) Page A-2

O’Shea, B., Clark, G. and Jankowski, J. 2006. A comparison of arsenic occurrence and

geochemistry in two coastal groundwater environments. 16th Annual V.M.

Goldschmidt Conference, 27 Aug – 1 Sept, Melbourne, Australia.

Book Contributions O’Shea, B. and Jankowski, J. 2006. Geogenic arsenic in an Australian sedimentary

aquifer: risk awareness for aquifers in Latin American countries. Proceedings from

the International Congress on Natural Arsenic in Groundwaters of Latin American

Countries, 20-22 June, Mexico City, Mexico. Balkema, Rotterdam.

Journal Articles O’Shea, B., Jankowski, J. and Sammut, J. (accepted July 2006) The Source of Naturally

Occurring Arsenic in a Coastal Sand Aquifer Of Eastern Australia. Science of the

Total Environment (special arsenic issue).

Arsenic in the Asia-Pacific Region Workshop Adelaide 2001

The Correlation between Arsenic, Pyrite and Marine Claysin a Coastal Aquifer

B. O’Shea, J.V.S. Smith and J. Jankowski

University of New South Wales, UNSW Sydney NSW 2052, AUSTRALIA

INTRODUCTIONPrevious studies of the Stuarts Point coastal sands aquifer, in northern New South Wales, have found elevated levels of arsenic in the coastal groundwater system (Smith et al., 2000; Smith et al., 2001). The two most important uses of this aquifer are; drinking water supply for the Stuarts Point township, and use of the groundwater for irrigation of local crops. In Australia, the National Health and Medical Research Council recommends an arsenic threshold of 7 g/L in drinking water (NHMRC, 1996). The Stuarts Point coastal aquifer has exceeded this limit by a factor of close to 50, with arsenic levels in the aquifer being above 330 g/L (Smith et al., 2001). Two possible hydrogeochemical processes that are responsible for arsenic release into groundwater are oxidation of arsenic enriched pyrite and pH influenced desorption from Al-OH sites on marine clays. The purpose of this current research is to complete extended investigations into the vertical solid phase distribution of arsenic in coastal sand aquifer matrices that could be responsible for the high levels already found in the local groundwater system.

RESULTS AND DISCUSSION Figure 1 shows the stratigraphic column constructed from data and samples obtained during drilling. A plot of arsenic concentration (mg/kg) with depth is also presented. Other dominant ions found in the nitric acid extract include Ca, Fe, S, Al, Mg, Na and K. Results of the XRD analysis showed that quartz was the dominate mineral, with illite the major clay mineral and pyrite present at depth. Arsenic concentrations ranged from 0.05–5.30 mg/kg in the acid extracts. Since world averages for arsenic in soil are around 5mg/kg, these results are not unusually high. The reason for studying the solid phase arsenic source lies in the fact that high levels of arsenic are known to be in the groundwater, and that high pH hinders arsenic adsorption onto clay minerals. Figure 1 shows the stratigraphy encountered several shell layers recording soil pH ranges of 8–9, and high Ca values in the acid extracts. These sediment characteristics indicate that arsenic may already be mobilised into the groundwater system. Two main processes for arsenic release into the Stuarts Point coastal aquifer are presented below.

Arsenic and Marine Clays Figure 1 shows arsenic increases in the profile in association with layers exhibiting higher clay contents (depths of 18m and 22.5-27m). The XRD analysis identified the major clay mineral as illite. During the Holocene when these clays were deposited, arsenic from the sea contributed to arsenic in the sediments. The arsenic is adsorbed onto Al-OH functional groups at particle edge sites of these illitic clays (Lin and Puls, 2000; Manning and Goldberg, 1997). A rise in pH due to the high content of shell material present in the aquifer is expected to have hindered adsorption and hence already mobilised arsenic into solution.

Arsenic and Pyrite During the Holocene, sea level rose and inundated the land. Sulfate in the seawater mixed with land sediments containing iron and organic matter, producing large amounts of iron

36


Figure 1. Stratigraphic column through the Stuarts Point coastal aquifer showing lithology, soil pH, arsenic concentrations (mg/kg) and other elements down the profile.

sulfides, namely pyrite (White et al., 1996). When exposed to air these sulfides oxidise to form sulfuric acid, hence the name acid sulfate soils (ASS). This reaction is shown in equation (1).

FeS2 + 15/4O2 + 7/2H2O Fe(OH)3 + 2SO42- + 4H+ (1)

The piezometer in this study is located within an ASS risk area due to its proximity to the coast and estimated age of deposition. Several studies have found arsenic is released from pyrite upon oxidation of ASS (Dudas, 1987; Gustaffson and Tin, 1994). XRD showed pyrite was found predominantly at depth, with only small amounts present closer to the surface. Figure 2 shows the XRD pattern for a sample at 27m, illustrating quartz, pyrite and illite peaks. As pyrite presence increases down the profile, so too does arsenic concentration. Those layers not containing illite, were found to contain pyrite, hence accounting for arsenic concentrations present in the sandy layers. Concentrations of Mg, Mn and Zn near the surface are low. This, combined with low arsenic and pyrite levels, indicates that pyrite oxidation has already taken place in these surface layers, hence releasing arsenic and other metals into the aqueous environment. pH is the only factor not correlating with the oxidation of ASS. pH would be expected to be much lower due to production of sulfuric acid (1). However, buffering by shell material and low amounts of pyrite present in this area would not lower pH dramatically. It is therefore concluded that the ASS risk for this piezometer is low. However, other areas of the Stuarts Point coastal aquifer have clay layers closer to the surface, and it is expected these areas would exhibit a much higher ASS risk, and therefore release more arsenic into the groundwater system.

37


Figure 2. XRD pattern for clayey sand sample at 27m depth, showing quartz, pyrite and illite peaks.

Quartz

Illitepyrite

Quartz

Quartz

pyrite

CONCLUSIONS Analysis of sediment samples in the Stuarts Point coastal sands aquifer indicates two possible hydrogeochemical processes are releasing arsenic into the groundwater system. The first is the oxidation of arsenic enriched pyrite contained within acid sulfate soils in the near surface sediments. The second process is desorption of arsenic from marine clays. Further study is in progress to analyse the full cycle of arsenic in this coastal environment.

ACKNOWLEDGEMENTS The authors would like to thank the Kempsey and Grafton offices of the New South Wales Department of Land and Water Conservation for their funding assistance.

REFERENCES Dudas, M.J. 1987. Accumulation of native arsenic in acid sulphate soils in Alberta. Can. J.

Soil Sci. 67, 317-331. Gustafsson, J.P. and Tin, N.T. 1994. Arsenic and selenium in some Vietnamese acid sulphate

soils. Science of the Total Environment. 151, 153-158. Lin, Z. and Puls, R.W. 2000. Adsorption, desorption and oxidation of arsenic affected by clay

minerals and aging process. Env. Geol. 39, 753-759. Manning, B.A. and Goldberg, S. 1997. Adsorption and stability of arsenic (III) at the clay

mineral-water interface. Environ. Sci. Technol. 31, 2005-2011. National Health and Medical Research Council & Agriculture and Resources Management

Council of Australia and New Zealand (NHMRC) 1996. Australian Drinking Water Guidelines. National Water Quality Management Strategy, Canberra.

Smith, J.V.S., Jankowski, J. and Sammut, J. 2000. Arsenic and acidity in a groundwater system affected by acid sulphate soils, Stuarts Point, Australia. In: O. Sililo et al. (eds) Proc. XXX Congress IAH, Cape Town, South Africa, 26 Nov – 1 Dec, 2000: 615-621. Rotterdam:Balkema.

Smith, J.V.S., Jankowski, J. and Sammut, J. 2001. Vertical distribution of As(III) and As(V) in a coastal sandy aquifer: Marine clays, a potential reservoir for arsenic in coastal groundwater systems. 10th International Symposium on Water-Rock Interactions, Villasimius, Italy, June 10-15, 2001. Balkema, Rotterdam.

White, I. And Melville, M.D. with Sammut, J., van Oploo, P., Wilson, P.B. and Yang, X. 1996. Acid Sulfate Soils – facing the challenges. Earth Foundation Australia Monograph 1. Millers Point, NSW.

38

ANTHROPOGENIC AND GEOMORPHIC FACTORS CONTRIBUTING TO ARSENIC DISTRIBUTION IN A COASTAL AQUIFER, NEW SOUTH WALES, AUSTRALIA

Bethany O’Shea and Jerzy Jankowski UNSW Groundwater Centre, School of Biological, Earth and Environmental Sciences, University of New

South Wales, Sydney, NSW, 2052, Australia

Abstract. Elevated levels of dissolved arsenic have been identified in an Australian coastal aquifer. The distribution of arsenic is controlled by Quaternary sands and clays deposited during subsequent sea level changes and fluvial depositional conditions. A number of processes are releasing arsenic into the aquifer, possibly the reductive dissolution of iron oxyhydroxides and the oxidation of arsenic enriched pyrite along flowpath. Anthropogenic use of the groundwater may be contributing to arsenic release into the aquifer, which can then enter environmental pathways such as plants/crops and organisms dependent upon the groundwater. The occurrence of arsenic in such coastal environments has management implications for these often heavily populated areas.

Keywords: arsenic, coastal aquifer, Quaternary geomorphology, geochemistry

INTRODUCTION

Coastal aquifers are often heavily relied upon for their groundwater resources. Exploitation of the groundwater can degrade both the quantity and quality of this natural resource. However, anthropogenic use of coastal groundwater is not the only factor contributing to detrimental water quality. Elevated levels of naturally occurring dissolved arsenic occur in groundwater systems throughout the world. The most notable occurrences include Bangladesh and West Bengal (Kinniburgh and Smedley, 2001), Argentina (Smedley et al., 2002), Taiwan (Chen et al., 1995), Chile (Sancha and Castro, 2001) and many parts of the USA (Welch et al., 2000). Often these areas are using the groundwater for domestic consumption, even though the concentrations of arsenic far exceed water quality guidelines.

Arsenic is a known human carcinogen. Exposure to this toxic element has been linked to skin cancer, keratosis and Blackfoot Disease (a type of gangrene) (Tseng, 1977). Other detrimental health effects include severe skin lesions, vascular and neurological effects, ulceration and even death (Thornton, 1996). Based on studies linking arsenic to these health problems, water quality criteria for arsenic is being reviewed around the world. The current WHO guideline stands at 10 g/L (WHO, 1994), while the US EPA is presently assessing its standing criteria level of 50 g/L and the possibility of lowering this level to 10 g/L. The maximum permissible limit for dissolved arsenic in Australian drinking water is 7 g/L (NHMRC, 1996).

Naturally occurring arsenic in groundwater is derived from a number of different geological sources. These include:

arsenic released from mining activities, particularly the oxidation of arsenopyrite, a sulphide mineral commonly associated with ores of gold, tin, silver and zinc (Welch et al., 1988);

arsenic associated with geothermal waters and hot springs (Welch et al., 1988);

the oxidation of arsenic enriched pyrite (a common component of acid sulfate soils) by dissolved oxygen (Dudas, 1984);

pH influenced desorption of arsenic species from surface sites of clays (Manning and Goldberg, 1997) and iron/aluminium/manganese oxyhydroxides (Ghosh and Yuan, 1987; Raven et al., 1998; Driehaus et al., 1995);

arsenic released via reductive dissolution of oxyhydroxides (Ravenscroft et al., 2001); and

release of arsenic into solution by competition with phosphate for the same adsorption sites on mineral surfaces (Acharyya et al., 2000).

This study looks at the distribution of arsenic in a coastal aquifer, in order to determine the geochemical processes responsible for its mobilisation into the groundwater system. The most likely processes controlling arsenic geochemistry at this site is due to the presence of Quaternary sediments of marine, fluviatile and deltaic depositional origin. Analysis of anthropogenic activities, geomorphology, and arsenic occurrence in sediments and groundwater are presented to

determine the geochemical cycle of arsenic in this Australian coastal aquifer. The data presented herein forms part of a continuing study in the area.

METHODOLOGY

Three multilevel piezometers were installed in the coastal aquifer to depths of 28m (BOS1), 30m (BOS2) and 22m (BOS3). Drilling was performed by a hollow stem auger without the use of any drilling muds. A sediment sample was taken by a split spoon sampler every 1.5m. These sediments were oven dried, ground and sieved through a 250 m sieve. An acid extract was then conducted with 10% nitric acid and the supernatant analysed by ICP-AES. XRD was performed on all sediment samples from BOS1.

The multilevel piezometers were constructed so that a groundwater sample could be taken every metre. Samples were collected from BOS1 and BOS3 from the water table level to the bottom of the piezometer. Sample collection occurred after

general parameters (pH, Eh, EC, dissolved oxygen and temperature) had stabilised. Major cations and trace elements – including arsenic – were analysed by ICP-AES, while unstable parameters such as CO2, HCO3

-, S2-, Fe2+, NO3-, NH4

+ and PO43- were

analysed in the field. Cl- was determined by argentometric titration with silver nitrate in the laboratory. A charge balance error calculation was performed in order to check the electroneutrality of the analysed groundwater samples.

ENVIRONMENTAL SETTING

Stuarts Point is located approximately 400km north of Sydney, on the New South Wales Mid-North coast (Figure 1). The area is a mix between resident farmers and beachside holiday makers. Farming includes avocadoes, potatoes, peaches, macadamia nuts and strawberries. Most farmers rely on the groundwater to irrigate their crops.

Figure 1. Location Map and surface geomorphology of the Stuarts Point aquifer. The inner Pleistocene sand barrier is denoted by ‘cy’ (blue) and the outer Holocene sand barrier by ‘sp’ (yellow). BOS1, 2 and 3 are the locations of the three

multilevel piezometers used for this study. This map is adapted from Eddie (2000).

Regional geology is comprised of sandstones, siltstones, mudstones and conglomerates of the Kempsey and Pee Dee beds(Leitch, 1972). In the north of the catchment lies the Yarrahapinni Mountain, a coastal granitoid consisting of hornblende-biotite monzogranite and alaskite, with contact aureoles of foliated sandstones, phyllites and sandy slate (Gilligan et al., 1992).

Groundwater for town and farm use is extracted from the coastal sands aquifer shown in Figure 1 as the Holocene sand barrier. The other main geomorphic unit in the study area is the inner Pleistocene sand barrier.

Both units were deposited during the late Quaternary period. The landward edge of the inner Pleistocene barrier represents the pre-Pleistocene shoreline. This former shoreline is shown by the dashed line in Figure 2.

Figure 2. Geomorphological map of part of the Mid-North coast showing inner and outer barriers and

former shorelines (Hails and Hoyt, 1968).

The dual barrier system initially developed due to submergence of dune or beach ridges adjacent to pre existing shorelines (Hoyt, 1966). As sea level slowly rises, a shallow lagoon is formed behind the raised beach ridge. On the New South Wales Mid-North coast progradation of these two barriers occurred by onshore movement of material at stillstands during submergences (Hails, 1968). These barriers are separated by inter barrier lagoonal swamps (such as the Yarrahapinni Wetlands) that represent a time interval between deposition of the inner and outer barriers at Stuarts Point. The headlands at Grassy Head and Smoky Cape were once offshore islands (ie, bedrock) which have since been joined to the coastline by deposition of this dual barrier system.

The inner Pleistocene barrier sands are quartzose dominant and often contain coffee rock. The outer Holocene barrier is also quartzose and contains abundant shell material. Shells can be naturally accumulated, or form part of aboriginal shell midden complexes located throughout the area (Sullivan and Hughes, 1978). There is evidence of barrier breaching by the Macleay River, as can be seen in the cross section of Figure 3, and the deflection of the Macleay River northwards within the outer barrier.

BOS3 is composed mainly of sand until a sandy clay unit was penetrated 6m above the point of drill rig refusal. It is assumed that bedrock was penetrated at 24m depth, and clay located above this is weathered rock. However, BOS1 and 2 are significantly different in terms of sedimentology. Deltaic, fluvial and estuarine sediments underlie the sands of the outer barrier indicating a change in depositional conditions for these two bores. The inner barrier is possibly lain underneath these estuarine and fluvial sediments, with both bores terminating in clays. The bedrock topography in the study area is not known, therefore geomorphic reconstruction below 30m depth in the outer barrier is difficult. Old drill records indicate bedrock to be lower than 60m in the proximity of BOS2 and BOS1. Therefore, it is assumed these unconsolidated sands and clays extend much deeper than 24m where bedrock was encountered in BOS3.

GROUNDWATER CHEMISTRY

General aquifer chemistry

The cross section in Figure 3 graphically shows the major water classifications (listed in Table 1) for selected depths in the two multilevels

sampled, BOS1 and BOS3. It can be seen that BOS3 is predominantly a Ca-HCO3 type groundwater influenced by the quartzose sand of the outer barrier. However, further down the flowpath at BOS1, there are several different groundwater types present (Ca-HCO3, Ca-Na-

HCO3-Cl and Ca-Mg-HCO3). Water classification is thus highly dependent on groundwater flow through the different sedimentological units. These sedimentological units are the result of changes in depositional environments during the late Quaternary.

Figure 3. Cross section from BOS3 to BOS1 showing geomorphology and potential depositional environments, groundwater flow direction, and major groundwater chemistry.

Table 1. Major water classification types for selected depths in BOS3 and BOS1. Ions were included in the classification if their concentration exceeded 20% of the total cations and anions respectively, calculated in meq/L.

Depth BOS3 BOS16m below surface Ca-HCO3 Ca-Na-HCO3

12m below surface Ca-Na-HCO3-Cl Ca-HCO3

18m below surface Ca-HCO3 Ca-Mg-HCO3

24m below surface No sample Na-Cl-HCO3

Due to the abundance of shell material in the aquifer sediments, dissolution of calcite is one of the major geochemical processes occurring. This accounts for Ca-HCO3 dominating as a major water type, as well as contributing small amounts of Mg2+ and Sr2+ (Figure 4).

Cl- concentrations are low and exhibit a good correlation with Na+ as shown in Figure 5. The occurrence of Cl- is thus most likely due to salt being deposited during coastal rainfall events and leaching into the groundwater. EC was also low (300 – 1,500 S/cm) considering the proximity of

the piezometers to the coast and the potential threat of saltwater intrusion. The higher EC values were consistent with the clay layers found at depth in BOS1. Generally groundwaters are slightly reducing and have a neutral pH. BOS1 was more reducing than BOS3 (-100 to –150mV).

0 1 2 3 4 5 6 7 8 9 10Cl (mmol/L)

0

1

2

3

4

5

6

7

8

9

Na

(mm

ol/L

)

LegendBOS3BOS1

Na:Cl correlation line

In BOS1, clay layers contribute to increases in Mg2+, Al3+ and Na+, thereby changing the dominance of the Ca-HCO3 waters as groundwater flow encounters these estuarine clay units.

The Stuarts Point coastal aquifer is prone to acid sulfate soil development due to its proximity to the coast and depositional history.

However, shallow sediments recovered during drilling did not exhibit characteristics of acid sulfate soils. Grey clays were found at depth down the profile, with pyrite being identified at 27m in BOS1. A plot of total Fe versus SO4

2-

(Figure 6) shows a slight increase in SO42- in

BOS1 compared to very low levels in BOS3. The oxidation of pyrite can occur by the equation shown below, or by nitrate acting as an oxidant.

FeS2-As + 15/4O2 + 7/2H2O Fe(OH)3 + 2SO4

2- + 4H+ + As

Both processes release Fe and SO42- into solution.

The increased SO42- in BOS1 may be coming

from pyrite oxidation somewhere along the flowpath between BOS3 and BOS1.

Figure 4 Ca+Mg+Sr versus HCO3 showing calcite dissolution is occurring in the aquifer.

Figure 5 Correlation between Na and Cl..

0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9SO4 (mmol/L)

0

0.01

0.02

0.03

0.04

0.05

0.06

0.07

0.08Fe

Tot

(mm

ol/L

)

LegendBOS3BOS1

BOS1 exhibits a better correlation than BOS3

Figure 6 Total Fe versus SO42-. SO4

2- may be derived from pyrite oxidation in BOS1, and from coastal rainfall.

0 1 2 3 4 5 6 7HCO3 (mmol/L)

0

0.5

1

1.5

2

2.5

3

Ca

+M

g+

Sr(m

mol

/L)

LegendBOS3BOS1

Calcite dissolution line

Arsenic geochemistry

Figures 7 and 8 show the vertical distribution of arsenic in each piezometer sampled. It can clearly be seen that BOS3 has much lower levels of arsenic and many that fall below the drinking water limit of 7 g/L. There is a slight increase in arsenic towards the bottom of the piezometer where bedrock was encountered. The vertical plot for BOS1 however shows levels of arsenic up to tenfold higher than the Australian drinking water limit. Arsenic distribution down the profile is erratic, and no good correlations exist with any one chemical reaction.

A slight correlation exists between Fe(II) and HCO3

- (Figure 9), indicating plausible dissolution of iron oxyhydroxides. The reductive dissolution of iron oxyhydroxides present as

Figure 7. Vertical arsenic profile of BOS3 with accompanying geological log.

Figure 8. Vertical arsenic profile of BOS1 with accompanying geological log.

Figure 9. Fe(II) versus HCO3- showing a slight

correlation between the two ions and the possibility of reductive dissolution of iron oxyhydroxides occurring.

Figure 10. A plot of As versus Eh showing the higher As concentrations occurring in the reducing Eh range of -100 to –150 mV.

surface coatings on grains can release arsenic into the groundwater. Figure 10 also supports this process. It shows As versus Eh. The higher levels of arsenic reported were between –100 to –150 mV therefore slightly reducing and potentially suitable for reductive dissolution to occur. There is also a possibility of arsenic being released via the oxidation of pyrite which may be a contributing reaction process between the two piezometers.

To determine the exact chemical reactions responsible for the distribution of arsenic in this coastal aquifer is a difficult task. A number of sources are governing the arsenic geochemical cycle. Since there is little arsenic in BOS3 which terminates in bedrock, the surrounding geology is not predicted as the major source of arsenic into this coastal aquifer. The highest concentrations

have been located downgradient, in the heterogenous sediments of the late Quaternary period. It is these sediments that are likely to be the dominant control over arsenic distribution in the aquifer.

2.5 3 3.5 4 4.5 5 5.5 6 6.5HCO3 (mmol/L)

0

0.01

0.02

0.03

0.04

0.05

0.06

Fe(I

I)(m

mol

/L)

LegendBOS3BOS1

Initially, sediment analyses (nitric acid extract) reported the highest solid phase arsenic to be found in the clay units of all three piezometers. The clay units are thus suspected of adsorbing the dissolved arsenic as it is released. Arsenic is adsorbed onto Al-OH functional groups at particle edge sites of these clays.

ANTHROPOGENIC FACTORS AND THE DISTRIBUTION OF ARSENIC

No anthropogenic source of arsenic has been uncovered in the Stuarts Point coastal aquifer, but this does not mean that arsenic is not being influenced by human activities. Coastal aquifers are delicate ecosystems, and overpopulation of these aquifers can cause changes to the natural equilibrium state of the groundwater resource. Extraction of groundwater for domestic and irrigation supply may be contributing to increased mobilisation of arsenic through the oxidation of arsenic enriched pyrite.

-150 -100 -50 0 50 100 150 200 250 300 350Eh (mV)

0

10

20

30

40

50

60

70

As(

g/L

)

LegendBOS3BOS1

The addition of farming chemicals such as phosphate has been confirmed in the region. Phosphate will compete for the same adsorption sites as arsenic. Studies of the soil in avocado fields of the study area have revealed extremely high levels of total phosphorous in solid phase. It is likely that if arsenic was previously present in solid phase, phosphorous, if present as phosphate, may have taken its adsorption site and released arsenic into solution for uptake into crops and/or leaching towards the water table. Therefore this region may not only be exposed to excessive levels of dissolved arsenic in their groundwater drinking supply, but also as an accumulant in the crops they are producing.

IMPLICATIONS FOR GROUNDWATER MANAGEMENT IN COASTAL ZONES

The original source of arsenic in the Stuarts Point coastal aquifer is thought to be derived from seawater when sediments were deposited. If this is proved to be correct, then coastal zones of Australia which have been subjected to similar depositional histories may be prone to elevated levels of arsenic in their coastal aquifers. Studies conducted by Hails (1968), Hails and Hoyt (1968) and Hoyt (1966) surrounding the Stuarts Point

aquifer discuss sea level change and depositional histories similar to most of the east coast of Australia, and also other parts of the world. Since coastal areas are often heavily populated, withdrawal of groundwater may contribute to increased arsenic in the aquifer via changes to the aquifer’s natural equilibrium state. Groundwaters discharging to nearby ecosystems such as wetlands and estuaries may contain high arsenic concentrations, which can in turn affect organisms living in these ecosystems. Both natural and farmed plants have the potential to be influenced by arsenic uptake through their root systems. A full understanding of the arsenic geochemical cycle is needed in order to assess the implications of this element on coastal environments.

CONCLUSION

Elevated levels of dissolved arsenic in the Stuarts Point coastal aquifer and sedimentology have been identified. The late Quaternary geomorphic history of the area is thought to be governing the cycle of arsenic in solid and aqueous phase. A number of chemical processes are contributing to the release of arsenic into the aquifer, the most dominant process being the dissolution of iron oxyhydroxides and subsequent release of arsenic into solution, aided by minor oxidation of arsenic enriched pyrite. Estuarine clays are thought to be adsorbing this dissolved arsenic onto their surface sites along the flowpath.

It is possible that arsenic may be present in other coastal aquifers of similar depositional history. This can have management implications for coastal populations using the groundwater as a resource.

REFERENCES

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Chen SL, Yeh SJ, Yang MH, Lin TH. 1995. Trace element concentration and arsenic speciation in the well water of a Taiwan area with endemic Blackfoot disease. Biol. Trace Elem. Res. 48, 263-274.

Driehaus W, Seith R, Jekel M. 1995. Oxidation of arsenate(III) with manganese oxides in water-treatment. Water Res. 29, 297-305.

Dudas MJ. 1984. Enriched levels of arsenic in post-active acid sulfate soils in Alberta. Soil Sci. Soc. Am. J. 48, 1451-1452.

Eddie MW. 2000 Soil Landscapes of the Macksville & Nambucca 1:100 000 Sheets,Department of Land and Water Conservation, Sydney.

Gilligan LB, Brownlow JW,. Cameron RG and Henley HF.1992. Dorrigo – Coffs Harbour 1:250 000 Metallogenic Map SH/56-10, SH/56-11: Metallogenic Study and Mineral Deposit Data Sheets. New South Wales Geological Survey, Sydney.

Ghosh MM, and Yuan JR. 1987. Adsorption of inorganic arsenic and organoarsenicals on hydrous oxides. Environ. Prog. 6, 150-157.

Hails JR. 1968. The late Quaternary history of part of the mid-north coast, New South Wales, Australia. Trans. Inst. Br. Geogr., 44, 133-149.

Hails JR, and Hoyt JR.1968. Barrier development on submerged coasts; problems of sea level changes from a study of the coastal plain of Georgia, U.S.A., and parts of the east Australian coast. Z. Geomorph.

Hoyt JH. 1966. Barrier island formation (abstract) Geol. Soc. Am. Programme for Annual Meeting, 99.

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Leitch EC. 1972. The Geological development of the Bellinger-Macleay Region- a study of the tectonics of the New England Fold Belt. PhD Thesis, University of New England.

Manning BA and Goldberg S. 1997. Adsorption and stability of arsenic(III) at the clay mineral-water interface. Environmental Science & Technology, 31(7), 2005-2011.

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Ravenscroft P, McArthur JM and Hoque BA. 2001. Geochemical and palaeohydrological controls on pollution of groundwater by arsenic. In: Chapell WR, Abernathy CO, Calderon RL. (eds) Arsenic Exposure and Health Effects IV. Elsevier, Amsterdam, 20p.

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NATURAL ARSENIC IN A COASTAL AQUIFER: IMPLICATIONS FOR WATER QUALITY MANAGEMENT IN AUSTRALIAN

COASTAL ZONES

Bethany O’Shea1and Jerzy Jankowski21PhD Research Student, UNSW Groundwater Centre, Sydney, NSW, Australia 2052

2Senior Lecturer, UNSW Groundwater Centre, Sydney, NSW, Australia 2052

Abstract

Arsenic concentrations in excess of Australian Drinking Water Guidelines have been found in a coastal aquifer at Stuarts Point on the Mid-North coast of New South Wales, Australia. Groundwater from this aquifer is used for a variety of purposes including domestic and town water supplies, and irrigation of local crops. Sediment and groundwater were analysed from a 27m deep multilevel piezometer. Results of this study indicate that arsenic is derived from several sources. The oxidation of pyrite present as coatings on sediment grains by both dissolved oxygen and nitrate is suggested by water chemistry and soil extract results. Reductive dissolution of iron oxyhydroxide coatings could also contribute arsenic to the system, but these coatings have not yet been identified in the system. pH influenced adsorption/desorption of arsenic on surface sites of clays is probable, and is being investigated further. Arsenic and element profiles with depth indicate a combination of various sources are contributing arsenic to the groundwater system at different depths in the profile. Arsenic in this aquifer is derived from a natural source. Other coastal areas with similar depositional histories should be aware of the possibility of elevated arsenic in their groundwater systems, and incorporate this into their management plans.

Key Words: arsenic, coastal aquifers, pyrite, desorption, iron oxyhydroxides

Introduction

Elevated levels of dissolved arsenic have been found in groundwater in various parts of the world. The most notable occurrences include West Bengal and Bangladesh, Hungary, Taiwan, China, Chile, Argentina and many parts of the USA. Millions of people are being exposed to arsenic by consumption of groundwater used for domestic supplies. Exposure to arsenic can cause skin lesions, kerotosis, cancer, ulceration, gangrene and ultimately death.

Natural arsenic in groundwater is derived from a number of different geological sources. These include:

arsenic released from mining activities, particularly the oxidation of arsenopyrite, a sulphide mineral commonly associated with ores of gold, tin, silver and zinc

arsenic associated with geothermal waters and hot springs

the oxidation of arsenic enriched pyrite (a common component of acid sulfate soils) by dissolved oxygen or other oxidants, such as nitrate

pH influenced desorption of arsenic species from surface sites of clays and iron and aluminium oxyhydroxides

arsenic released via reductive dissolution of iron oxyhydroxides, and

release of arsenic into solution by competition with phosphate and bicarbonate for the same adsorption sites on mineral surfaces.

The purpose of this study was to determine possible sources and release mechanisms of natural arsenic in an Australian coastal aquifer. Once these sources and mechanisms have been identified an assessment of arsenic occurrences in other coastal zones can be made.

Study Location

Stuarts Point is located on the Mid-North coast of New South Wales, approximately 400km from Sydney. Previous studies of the Stuarts Point coastal aquifer have recorded levels of dissolved arsenic in excess of 330 g/L (Smith et al., 2001). Arsenic has been found throughout the entire catchment, at different depths and in different lithological units (Smith et al., 2000). The limit imposed on drinking water in Australia is just 7 g/L (NHMRC, 1996). This means the levels found in the Stuarts Point aquifer are up to 45 times higher than the limit assigned to drinking water. Groundwater from this aquifer is used to supply town water for both domestic consumption and irrigation of local crops.

Geomorphology

Figure 1 shows the main units of the Stuarts Point unconsolidated coastal sediments.

Figure 1. Lithology of the Stuarts Point unconsolidated sediments (a) Holocene

barrier sands (b) alluvial swamp deposits (c) estuarine clay (d) Pleistocene barrier sands (Smith et al., 2000). The red dot indicates location of the multilevel piezometer installed for this study.

The aquifer is comprised of Pleistocene and Holocene barrier sands deposited by wind and wave action, estuarine muds and clays, and alluvial swamp deposits.

Like many coastlines, the Stuarts Point sediments reflect the interaction between coastal, estuarine and fluvial depositional processes. During the Holocene, the east coast of Australia was subjected to a sea level transgression. Sea level rose and inundated the land, causing seawater to mix with sediments. Arsenic in the seawater is thought to have been incorporated into the sediment during this time.

The unconsolidated sediments overlie sandstones, siltstones, mudstones and conglomerates of the Kempsey and Dee Pee beds (Leitch, 1972). In the north of the catchment lies the Yarrahapinni Mountain, a coastal granitoid consisting of hornblende-biotite monzogranite and alaskite, with contact aureoles of foliated sandstones, phyllites and sandy slate (Gilligan et al., 1992).

Mineralogy and Sediment Chemistry a Sediment samples were collected every 1.5

metres during drilling of a multilevel piezometer. The piezometer for this study was drilled in the Holocene barrier sand unit, the exact location is shown in Figure 1. Figure 2 shows the stratigraphy encountered. A nitric acid extract was carried out on all samples to determine extractable element concentrations. Figure 2 shows sediment arsenic concentrations (mg/kg) down the profile.

b

c

Sediment arsenic concentrations ranged from 0.05 – 5.30 mg/kg in the acid extracts. Since world averages for arsenic in soil are around 5 mg/kg, these results are not unusually high. However high arsenic groundwaters do not usually appear to be directly related to areas of high arsenic concentrations in the aquifer sediments (BGS, 2001).

d

Other dominant elements found in the nitric acid extract include Ca, Fe, S, Al, Mg, Na and K. Shell material is responsible for Ca values, pyrite contributes to Fe and S concentrations, and Al, Mg, Na and K are most likely found in clays down the profile.

Sediments were analysed for mineralogical composition by X-Ray Diffraction (XRD). Since the sediments are dominated by sand, quartz was found to be the dominant mineral, with small amounts of feldspar, pyrite and the clay mineral illite, varying in depth down the profile.

Figure 2. Stratigraphaphic column through the Stuarts Point coastal aquifer showing arsenic concentrations in both sediments and groundwater.

Controls on arsenic distribution in sedimentsSediments in the study area are classified as having a high acid sulfate soils risk due to their depositional history and proximity to the coast. Studies by Dudas (1984, 1987) have identified accumulation of native arsenic in acid sulfate soils in Alberta, while in Vietnam arsenic has been linked to acid sulfate soils by Gustafsson and Tin (1994).

Pyrite, a common mineral in acid sulfate soils, can coprecipitate arsenic into its structure. When

pyrite is oxidised the arsenic is released into solution. This process is illustrated in equation (1):

FeS2-As + 15/4O2 + 7/2H2OFe(OH)3 + 2SO4

2- + 4H+ + Dissolved As (1)

A small amount of pyrite was identified by XRD at a depth of 27m. Arsenic was not significant at any other depth, so arsenic contained within mineralogical pyrite is most likely at 27m depth. However, results of the nitric acid extract on the sediments revealed high levels of both iron and sulfur throughout the entire profile. No other iron sulfide minerals were identified by XRD, but this process does not identify coatings on grains with great accuracy. It is possible that arsenic is coprecipitated with pyrite present as coatings on the sand grains in the aquifer.

Alternatively, grains may be coated by iron oxyhydroxides which release arsenic via reductive dissolution or pH influenced desorption processes. The use of a scanning electron microscope will be employed in further studies of the Stuarts Point aquifer sediments to help identify the nature of grain coatings.

From Figure 2 it can be seen that arsenic peaks in the sediments with lithological units containing clay. The XRD analysis identified the major clay mineral as illite. Studies by Khan et al. (2000) of arsenic source in the Bengal delta have also identified illite as the major clay mineral present in sediments of Holocene age. Arsenic can be adsorbed onto Al-OH functional groups at particle edge sites of these illitic clays (Lin and Puls, 2000; Manning and Goldberg, 1997).

Groundwater Chemistry

Groundwater was collected and analysed every metre from the top of the water table (6m at time of sampling) to 26m depth. General parameters such as pH, Electrical Conductivity (EC), redox potential (Eh), temperature and Dissolved Oxygen (DO) were analysed in the field. Other unstable parameters were also measured at the time of sampling and included CO2, HCO3

-, S2-, Fe2+,NH4

+, NO3- and PO4

3-. Major ions and trace elements (including arsenic) were analysed by ICP-AES.

Groundwater is highly influenced by sediment lithology. Water in the sand and shell layers to 17m depth were dominated by Ca-HCO3 ions. At 17 to 22m the water type was Na-Ca-Cl-HCO3indicating a mixture of sand, shell and clay contributions to groundwater composition. Finally, the deeper sample points exhibited a Na-Cl-HCO3

water type being controlled by marine clays deposited at depth.

Dominant geochemical processes in groundwaterThe dissolution of shell material is contributing Ca2+, Mg2+, Sr2+ and HCO3

- to the system. This is shown in Figure 3 in Appendix A. An excess of HCO3

- is found in 3 samples close to the surface and can be attributed to soil zone CO2 input. A correlation exists with Na+ and Cl-, indicating rainfall in the area is influenced by marine contributions, which is in turn contributing to aquifer chemistry. This is particularly noted close to the water table (6 to 9m depth) in a fine grained sand layer where Na+, SO4

2-, EC and NH4+ are in

excess. These elements increase again in marine deposited clay layers at depth. Due to the piezometer location being close to the coast, remnant seawater is likely to be slowly flushing out of these clays in the aquifer. Pyrite oxidation is contributing Fe and S species, and weathering processes are releasing minor amounts of Al3+,K+, silica and trace elements.

Controls on arsenic distribution in groundwaterAll major ion plots showed no direct correlation between arsenic concentrations and known release mechanisms into groundwater. For example, if pyrite oxidation is the only process releasing arsenic, then Fe2+ and SO4

2- should correlate, and arsenic should increase accordingly. Desorption of arsenic from clays or oxyhydroxide surfaces would show a correlation between arsenic and pH. Reductive dissolution of iron oxyhydroxides would show a correlation between FeTotal and HCO3

- as arsenic increased. It is therefore concluded that a combination of arsenic release processes are occurring in the Stuarts Point coastal aquifer. Examining arsenic and element changes with depth allows several processes to be identified at various depths down the profile.

Figure 4 shows a plot of arsenic, DO and nitrate with depth. At depths of 6-9m arsenic is decreasing in conjunction with decreases in DO. This could indicate DO is oxidising pyrite in these layers, which is then releasing arsenic into the groundwater system. However, DO does not correlate well with arsenic for the remainder of the profile. NO3

- can also oxidise pyrite as shown in equation (2):

10FeS2 + 30NO3- + 20H2O

10Fe(OH)3 + 15N2 + 15SO42- + 5H2SO4 (2)

however, this process removes NO3- from solution

while increasing arsenic, which cannot be seen in the profile. This process cannot be ruled out entirely because it is kinetically slow and so may still be occurring while NO3

- is increasing in solution. It is possible therefore that pyrite present as grain coatings is being oxidised by DO and NO3

- further down the profile, but there is no good correlation because other processes are beginning to contribute or remove arsenic in solution.

Figure 4. Plot of As, DO and NO3- with depth.

0.80 1.00 1.20 1.40 1.60 1.80DO (mg/L)

-26.00-25.00-24.00-23.00-22.00-21.00-20.00-19.00-18.00-17.00-16.00-15.00-14.00-13.00-12.00-11.00-10.00-9.00-8.00-7.00-6.00

Dep

th (m

bgs)

0.00 0.02 0.04 0.06 0.08As (mg/L)

0.00 0.40 0.80 1.20 1.60 2.00NO3 (mg/L)

DO

As

NO3

Figure 5 supports pyrite oxidation as an arsenic release mechanism from 6-16m depth. This is shown by similar patterns of SO4

2- and arsenic concentration fluctuations. When pyrite is being oxidised it releases SO4

2- and arsenic into solution at the same time.

From 13-18m there is also a good correlation between arsenic and Eh (Fig. 6). It is possible that arsenic release into solution may be favourable for some reactions at specific Eh ranges. The reductive dissolution of iron oxyhydroxides can contribute arsenic to the groundwater. This process is thought to be the main process

releasing extremely high levels of arsenic to groundwater in Bangladesh (BGS, 2001; Ravenscroft et al., 2001; Nickson et al., 2000).

However, there exists no good correlation between the products of iron oxyhydroxide reductive dissolution, which are FeTotal and HCO3

-.This is because Fe2+ is being released from pyrite oxidation, and HCO3

- is also being contributed by calcite dissolution, as shown in Figure 3 in Appendix A. It should be noted that there is an excess of HCO3

- on the calcite dissolution graph, indicating another source of HCO3

- input to the system, which could be attributable to reductive dissolution of iron oxyhydroxides on sediment grain coatings in the aquifer.

Figure 5. Plot of arsenic and sulphate with depth.

Sediment concentrations of arsenic revealed peaks associated with clay layers in the profile. Arsenic is adsorbed onto surface hydroxyl sites of these clays according to pH and arsenic species. Lin and Puls (2000) found that As(V) adsorption began to decrease at pH 7.5 while As(III)

adsorption increased at pH 7.5. The average pH of the groundwater from the clay layers was 7.5, so it is assumed that adsorption/desorption processes are contributing to arsenic release/decrease in the groundwater in these layers. Since arsenic in groundwater is decreasing in these clay zones (see Fig. 2) it is predicted that arsenic is being adsorbed onto clay surfaces at these locations. Further study is in progress to determine this mechanism.

Due to the complex nature of reactions occurring in this coastal aquifer, other geochemical reactions controlling the distribution of arsenic are currently being assessed.

-160.00 -140.00 -120.00 -100.00 -80.00Field Eh (mV)

-26.00-25.00-24.00-23.00-22.00-21.00-20.00-19.00-18.00-17.00-16.00-15.00-14.00-13.00-12.00-11.00-10.00-9.00-8.00-7.00-6.00

Dep

th (m

bgs)

0.00 0.02 0.04 0.06 0.08As (mg/L)

Eh

As

0.00 5.00 10.00 15.00 20.00 25.00 30.00Sulphate (mg/L)

-26.00-25.00-24.00-23.00-22.00-21.00-20.00-19.00-18.00-17.00-16.00-15.00-14.00-13.00-12.00-11.00-10.00

-9.00-8.00-7.00-6.00

Dep

th (m

bgs)

0.00 0.02 0.04 0.06 0.08As (mg/L)

As

SO4

Figure 6. Plot of arsenic and Eh with depth.

Management Implications for Australian Coastal Aquifers

The arsenic in the Stuarts Point coastal aquifer is a natural occurrence, deposited during sea level

rise in the Holocene. Similar sediments were deposited along the entire east coast of Australia during this time. It is possible then, that other coastal zones have elevated levels of natural arsenic in their groundwater systems. It is already known that arsenic can be associated with acid sulfate soils, and these soils are widespread in Eastern Australian coastlines. Fluctuations in water levels of aquifers due to extraction for irrigation, construction and drinking water can lead to oxidation of these sediments and release of arsenic. Introduction of phosphate to the system through anthropogenic activities could also release arsenic into groundwater. Levels of dissolved arsenic need to be measured in similar coastal zones and management practices to treat or control dissolved arsenic in coastal aquifers be put into place.

Conclusion

Natural arsenic was deposited in sediments at Stuarts Point during sea level rise in the Holocene. This arsenic is now being released into the groundwater system. There are several processes responsible for arsenic mobilisation in this coastal aquifer.

The oxidation of arsenic enriched pyrite by DO or NO3

-, reductive dissolution of iron oxyhydroxide coatings on sediment grains, and pH influenced desorption from surface sites of clays are thought to be the main processes controlling arsenic in this coastal aquifer.

Due to the complex nature of reactions occurring in the Stuarts Point groundwater system, further studies are in progress to determine the full cycle of arsenic in this environment.

Acknowledgements

The authors would like to thank the Grafton office of the New South Wales Department of Land and Water Conservation for their financial contribution to this study.

References

British Geological Survey (BGS) (2001) “Arsenic contamination of groundwater in Bangladesh. Volume 2: Final report”. BGS Technical Report WC/00/19.

Dudas, M.J. (1984) Enriched levels of arsenic in post-active acid sulfate soils in Alberta. Soil Sci. Soc. Am. J. 48: 1451-1452.

Dudas, M.J. (1987) Accumulation of native arsenic in acid sulfate soils in Alberta. Can. J. Soil Sci. 67: 317-331.

Eddie, M.W. (2000) Soil Landscapes of the Macksville & Nambucca 1:100 000 Sheets, Department of Land and Water Conservation, Sydney.

Gustafsson, J.P. and N.T. Tin (1994) Arsenic and selenium in some Vietnamese acid sulphate soils. Science of the Total Environment 151: 153-158.

Gilligan, L.B., J.W. Brownlow, R.G. Cameron & H.F. Henley (1992) Dorrigo – Coffs Harbour 1: 250 000 Metallogenic Map SH/56-10, SH/56-11: Metallogenic Study and Mineral Deposit Data Sheets. New South Wales Geological Survey, Sydney.

Khan, A.A., S.H. Akhter and S.M.M. Alam (2000) Evidence of Holocene trangression, dolomitization and the source of arsenic in the Bengal delta. Proceedings of the First International Conference on geotechnical, Geoenvironmental Engineering and Management in Arid Areas (GEO2000), November 4-7, 2000. A.A. Balkema Publishers, Rotterdam, Netherlands.

Leitch, E.C. (1972) The Geological development of the Bellinger-Macleay Region- a study of the tectonics of the New England Fold Belt. PhD.Thesis, University of New England.

Lin, Z. and R.W. Puls (2000) Adsorption, desorption and oxidation of arsenic affected by clay minerals and aging process. Env. Geol. 39,753-759.

Manning, B.A. and S. Goldberg (1997) Adsorption and stability of arsenic (III) at the clay mineral-water interface. Environ. Sci, Technol. 31, 2005-2011.

National Health and Medical Research Council & Agriculture and Resources Management Council of Australia and New Zealand (NHMRC) (1996) Australian Drinking Water Guidelines. National Water Quality Management Strategy, Canberra.

Nickson, R.T., J.M. McArthur, P.Ravenscroft, W.G. Burgess and K.M. Ahmed (2000) Mechanism of arsenic release to groundwater, Bangladesh and West Bengal. AppliedGeochemistry 15: 403-413.

Ravenscroft, P., J.M. McArthur and B.A. Hoque (2001) Geochemical and Palaeohydrological Controls on Pollution of Groundwater by arsenic. In: Chappell, W.R., C.O. Abernathy and R.L. Calderon (eds) Proceedings of the 4th

International Conference on Arsenic Exposure and Health Effects. Elsevier Science Ltd., Oxford.

Smith, J.V.S., J.Jankowski and J.Sammut (2000) Arsenic and acidity in a groundwater system affected by acid sulphate soils, Stuarts Point, Australia. In: O. Sililo et al (Editors), Groundwater:Past Achievements and Future Challenges,

Proceedings of the XXX Congress of the International Association of Hydrogeologists,Cape Town, South Africa, 26 November – 1 December 2000, Balkema, Rotterdam, 615-621.

Smith, J.V.S., J. Jankowski and J. Sammut (2001) Vertical distribution of As(III) and As(V) in a coastal sandy aquifer: Marine clays, a potential reservoir for arsenic in coastal groundwater systems. In: Proc. 10th International Symposium on Water-Rock Interactions, Villasimius, Italy, June 10-15, 2001. Balkema, Rotterdam, 1009-1012.

Appendix A

0.00

0.50

1.00

1.50

2.00

2.50

3.00

0.00 2.00 4.00 6.00 8.00HCO3 (m mol)

Ca+

Mg+

Sr (m

mol

)

1:1 calcite dissolution line

Figure 3. Calcite dissolution contributes Ca, Mg, Sr and HCO3 to the groundwater.

Authors Biographies

Bethany O’Shea is a PhD research student at the University of New South Wales Groundwater Centre. Her research involves determining the source, release and transport of natural arsenic in a coastal aquifer. Arsenic speciation and cycling in both sediments and groundwater is also a major component of her PhD project. Bethany has a First Class Honours degree in Earth and Environmental Science, and has worked for International Mining Consultants and the NSW Department of Mineral Resources during her undergraduate years. She is currently employed as hydrogeologist in the Sydney office of Egis Consulting Australia Pty Limited.

Postal Address: School of Geology, University of New South Wales, UNSW SYDNEY NSW 2052

E-mail: [email protected]

Jerzy Jankowski has a BSc in hydrology, an MSc in hydrology/hydrochemistry and a Ph.D. in hydrogeology/hydrogeochemistry, all from the University of Wroclaw, Poland. He has more than 20 years research experience in countries including Poland, Czechoslovakia, Bulgaria and Australia. He is author/co-author of over 100 research papers, chapters in books and conference proceedings. Areas of special research expertise include: water-rock interaction in different rock type environments, contamination of surface waters and groundwaters, chemical evolution of groundwaters in the arid zone, groundwater-sea water and groundwater-surface water interactions, dryland salinisation processes and hydrochemical and isotopic studies including environmental isotopes and hydrogeochemical modelling.

Postal Address: School of Geology, University of New South Wales, UNSW SYDNEY NSW 2052

E-mail: [email protected]

mailto:[email protected]


Irrigation Australia 2002 Conference, “Irrigation – Conservation or Conflict”, Sydney, 21-23 May 2002, B.G. Sutton (Editor), ISBN 0-9586424-5-1

IRRIGATION CONFLICTS IN A COASTAL AQUIFER – RISKS FROM NATURAL ARSENIC CYCLING

B. O’Shea and J. Jankowski UNSW Groundwater Centre, School of Biological, Earth and Environmental Sciences, University of New South

Wales, Sydney NSW 2052, Australia

ABSTRACT

Elevated levels of dissolved arsenic have been found in an Australian coastal aquifer used for irrigation of local crops. Arsenic is a carcinogen and therefore detrimental to human health. Groundwater with arsenic derived from natural sources is being extracted and applied to crops. Upon application to the surface environment, dissolved arsenic in this irrigation water has the potential to adsorb to soil particles, and be uptaken by crops, hence entering the food chain. Addition of nitrate and phosphate to the agricultural area can enhance the mobilisation of arsenic into the groundwater system. More research is needed to determine the extent of elevated arsenic in Australian aquifers. The effects of irrigation waters high in arsenic on the soil zone, and crop uptake of this element in Australian agricultural areas should be the subject of a detailed investigation.

Key words: arsenic, irrigation, adsorption, plant uptake

INTRODUCTION

Sustainable irrigation practices must obtain a balance between groundwater extraction and ecosystem equilibrium. Recent studies in an Australian coastal aquifer suggest that groundwater being used for irrigationcontains elevated levels of inorganic arsenic. Withdrawal of groundwater elevated in natural arsenic can have several effects on the surrounding environment:

the application of irrigation waters elevated in arsenic can concentrate this contaminant in the soil environment; uptake of arsenic by crops may lead to increased arsenic accumulation in the food chain; and groundwater extraction and other farming practices may increase the mobilisation of arsenic into these aquifers.

A variety of crops are grown in the coastal environment studied, including peaches, strawberries, avocadoes, macadamia nuts and potatoes. The uptake and accumulation of arsenic in these crops is not known, yet arsenic is a carcinogen and therefore detrimental to human health. This paper provides details of a study being conducted in a coastal irrigation area, and identifies the need for further research into the effects of arsenic cycling in the environment.

ARSENIC OCCURRENCES IN GROUNDWATER

Elevated levels of natural arsenic in groundwater have been found all over the world. In some countries, such as Bangladesh, West Bengal and parts of South America, levels are so high that they have been described as the worst mass poisoning in history (Anawar et al., 2002). Millions of people are being exposed to arsenic through drinking, cooking and bathing in arsenic contaminated groundwater. It is now thought that arsenic is entering the food chain via consumption of crops irrigated with this groundwater. Excessive exposure to arsenic can cause cancer, kerotosis, gangrene and ultimately death.

Natural arsenic in groundwater is derived from a number of different geological sources. These include: arsenic released from mining activities, particularly the oxidation of arsenopyrite, a sulphide mineral commonly associated with ores of gold, tin, silver and zinc; arsenic associated with geothermal waters and hot springs;

156


the oxidation of arsenic enriched pyrite (a common component of acid sulfate soils) by dissolved oxygen or other oxidants, such as nitrate; pH influenced desorption of arsenic species from surface sites of clays and iron, manganese and aluminium oxyhydroxides; arsenic released via reductive dissolution of oxyhydroxides; and release of arsenic into solution by competition with phosphate and bicarbonate for the same adsorption sites on mineral surfaces (O’Shea and Jankowski, 2002).

Arsenic concentrations in the coastal aquifer studied are up to 330 g/L (Smith et al., 2001), which exceed limits imposed on arsenic concentrations in both drinking water and irrigation water (Table 1). Arsenic may be originating from seawater accumulating in sediments deposited during the last 10,000 years. Research is continuing into this theory, but if proved correct, may have implications for all sediments deposited along the east coast of Australia which have undergone similar depositional conditions.

TABLE 1. Guidelines on arsenic concentration limits in drinking and irrigation water

Guideline NHMRC1 ANZECC2 WHO3 USEPA4

As concentration ( g/L) 7 100 10 10*

1 National Health and Medical Research Council (Australia) limit imposed on drinking water (NHMRC, 1996) 2 ANZECC (2000) long term trigger value for arsenic in irrigation water 3 World Health Organisation (1994) recommendations for drinking water 4 USEPA maximum contaminant level in drinking water (* under review)

Due to the many natural sources of arsenic, proximity of an aquifer to the coast is not a prerequisite for suspected contamination. Additional research conducted by the UNSW Groundwater Centre on an inland alluvial aquifer have also reported elevated concentrations of dissolved arsenic in groundwater being used extensively for irrigation (McLean and Jankowski, 2001). This illustrates the possibility that elevated levels of arsenic may be present throughout many Australian aquifers currently thought to be free of this toxic element.

IRRIGATION PRACTICES AND ARSENIC CONCERNS

The soil zone and shallow groundwater beneath irrigated lands gradually acquires a chemical signature influenced by the composition of applied irrigation water and any agricultural chemicals added (Schmidt, 1993). Thus, if arsenic is being added via irrigation water, it can become adsorbed onto soil particles (such as organic matter, clays and oxyhydroxides). It also has the potential to be uptaken by plants. This cycle of arsenic is shown in Figure 1. The use of arsenic contaminated groundwater for irrigating crops has led to elevated concentrations of arsenic in soils in Bangladesh and West Bengal. Normally soil arsenic concentrations range from 4–8 mg kg-1,but soils contaminated with arsenic can have concentrations of up to 60 mg kg-1 (Huq et al., 2001) resulting from applied irrigation water high in inorganic arsenic.

Effects on plants A study by Abedin et al. (2002) on the accumulation of arsenic in rice crops shows that elevated arsenic in irrigation water applied to rice plants decreased plant height, grain yield, the number of filled grains, grain weight, and root biomass. It also increased the amount of arsenic present in root, straw and rice husk. A similar study by Queirolo et al. (2000) in northern Chile showed arsenic accumulated in broadbeans, potatoes and maize after irrigation water high in arsenic (derived from a volcanic source) had been applied.

Root vegetables such as potatoes are most at risk to accumulate arsenic. Phytotoxicity in plants often results at concentrations of just a few mg kg-1 (McLaren et al., 2001).

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Other food chain effects The arsenic calamity of Bangladesh and West Bengal is now focussing on the impact of arsenic through the food chain. Apart from humans ingesting vegetables high in arsenic, cattle are also being exposed to elevated levels

As

As

AsAs in groundwater

extracted for irrigation

High As irrigation water applied to crops

As adsorbs to soil organic

Crops uptake

As infiltrates subsurface

Figure 1. The geochemical cycle of Arsenic (As) when present in irrigation waters.

through cattle feed, such as rice straw used as hay. This in turn provides another exposure pathway to humans, through the consumption of cattle meat.

Groundwater Extraction Although irrigation practices do not seem to be directly related to arsenic occurrences in groundwater, extraction may increase the rate of mobilisation of arsenic into the aquifer. Extraction of groundwater causes water levels to lower and sediments to become oxidised. If arsenic has been coprecipitated with pyrite (a theory currently being assessed for the coastal aquifer studied), it will be released upon oxidation via the following equation:

FeS2-As + 15/4O2 + 7/2H2O Fe(OH)3 + 2SO42- + 4H+ + Dissolved As (1)

X-Ray Diffraction of sediments from the aquifer have been conducted and identified pyrite in the matrix of some samples. Since most samples are predominantly sand, pyrite may also be present as coatings on these sand grains. The inland aquifer studied by McLean and Jankowski (2001) has pyrite identified present as grain coatings. Disturbing the natural equilibrium of these aquifers may therefore contribute to enhanced release of dissolved arsenic.

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Addition of agricultural chemicals Nitrate also has the potential to oxidise pyrite, according to equation 2 below:

10FeS2-As + 30NO3- + 20H2O 10Fe(OH)3 + 15N2 + 15SO4

2- + 5H2SO4 + Dissolved As (2)

The addition of ammonium nitrate has been confirmed in the coastal aquifer under consideration, however there is insufficient evidence to state that nitrate levels are directly responsible for the amount of arsenic present in dissolved phase. Preliminary groundwater results indicate that there are several processes governing arsenic occurrence in dissolved phase, with pyrite oxidation by nitrate being one of the processes possible (O’Shea and Jankowski, 2002).

Phosphate is another agricultural chemical that can influence the mobility of arsenic in solution. Phosphate and arsenate (the oxidised form of arsenic) have similar physicochemical characteristics, and compete directly for the same adsorption sites on mineral surfaces. When the two species are present at equal concentrations, phosphate will outcompete arsenate for adsorption sites (Abedin et al., 2002). Thus addition of phosphate to the soil environment can re-mobilise arsenic into solution, which will then leach down towards the water table, or increase availability of arsenic in the soil water to be uptaken by plants.

Selenium in the USA Another ion with similar properties to arsenate, is the oxidised form of selenium. In the United States, the National Irrigation Water Quality Program was set up in order to determine whether problems of selenium in irrigation drainwater were adversely affecting the health of humans, fish and wildlife. The source of selenium, like arsenic, was determined to be natural, accumulating in sediments of Cretaceous age from large amounts of volcanism during this period. Selenium has been commonly found at elevated concentrations in water, bottom sediment and biota (Presser et al., 1994; Feltz et al., 1991). In several areas, evaporative concentration of applied irrigation water and drainage of this water to wetlands, canals, streams, and lakes appeared to be responsible for elevated concentrations of selenium (Engberg and Sylvester, 1993).

These studies show the significance of the effects toxic elements can have when cycled through the ecosystem. Arsenic has the potential to effect humans and biota in a similar way to selenium, through the mismanagement of irrigation waters elevated in dissolved arsenic.

RECOMMENDATIONS FOR FURTHER RESEARCH

The recent discovery of high arsenic concentrations in groundwaters of West Bengal and Bangladesh have prompted authorities worldwide to re-evaluate the levels of arsenic in groundwater used for irrigation and drinking water supplies. In Australia, it is unknown as to how many aquifers contain elevated levels of arsenic. Those that do contain arsenic, need to be analysed in detail to assess the cycle of arsenic in solid and aqueous phase. Speciation in waters, sediments and biota would provide details not only on occurrence of different arsenic species, but also on toxicity of this element. Inorganic forms of arsenic are much more toxic than organic forms.

Where groundwater high in arsenic is being used for irrigation, studies need to be conducted to assess the influence of extraction, nitrate and phosphate addition on arsenic cycling. Uptake by crops needs to be further researched, and regular testing for arsenic levels in crops (particularly root vegetables) should be implemented. Treatment of irrigation water prior to application would help to prevent accumulation of arsenic in the food chain.

CONCLUSION

Elevated levels of arsenic, a known carcinogen, have been found in a coastal and inland aquifer of Australia. Both aquifers are extensively used for irrigation of local crops. The source of arsenic is expected to be natural, and hence could be widespread over many parts of Australia with similar geomorphic/geologic depositional

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histories. Irrigation water high in arsenic which is applied to soils and crops can be cycled into these surface environments. Arsenic has the potential to adsorb onto soil particles, and be uptaken by crops, therefore entering the food chain.

This paper outlines the need for extensive research to be undertaken by the irrigation industry, to determine the impacts of natural arsenic contamination on water, soil and crops in Australian agricultural areas.

REFERENCES

Abedin, M.D.J., Cresser, M.S., Meharg, A.A., Feldmann, J. and Cotter-Howells, J. (2002) Arsenic accumulation and metabolism in rice (Oryza Sativa L.), Environmental Science & Technology, 36, 962-968.

Anawar, H.M., Akai, J., Mostofa, K.M.G., Safiullah, S. and Tareq, S.M. (2002) Arsenic poisoning in groundwater: Health risk and geochemical sources in Bangladesh, Environment International, 27, 597-604.

ANZECC, (2000) Australian & New Zealand Guidelines for Fresh and Marine Water Quality, Volume 3 Primary Industries – Rationale and Background. Australia & New Zealand Environment & Conservation Council.

Engberg, R.A. and Sylvester, M.A. (1993) Concentrations, distribution, and sources of selenium from irrigated lands in Western United States, Journal of Irrigation and Drainage Engineering, 119, 522-536.

Feltz, H.R., Sylvester, M.A. and Engberg, R.A. (1991) Reconnaissance investigations of the effects of irrigation drainage on water quality, bottom sediment, and biota in the Western United States, in Mallard, G.E. and Aronson, D.A. [eds] Proceedings of the U.S. Geological Survey Toxic Substance Hydrology Program, Monterey, Calif., March 11-15, 1991: U.S. Geological Survey Water-Resources Investigations Report 91-4034, 319-323.

Huq, S.M.I., Ahmed, K.M., Sultana, N. and Naidu, R. (2001) Extensive arsenic contamination in groundwater and soils of Bangladesh, Arsenic in the Asia-Pacific Region Workshop Adelaide, November 20-23, 2001.

McLaren, R.G., Megharaj, M. and Naidu, R. (2001) Fate of arsenic in the soil environment, Arsenic in the Asia-Pacific Region Workshop Adelaide, November 20-23, 2001.

McLean, W. and Jankowski, J. (2001) The occurrence of arsenic in an alluvial aquifer system, Northern New South Wales, Arsenic in the Asia-Pacific Region Workshop Adelaide, November 20-23, 2001.

National Health and Medical Research Council & Agriculture and Resources Management Council of Australia and New Zealand (NHMRC) (1996) Australian Drinking Water Guidelines, National Water Quality Management Strategy, Canberra.

O’Shea, B. and Jankowski, J. (2002) Natural arsenic in a coastal aquifer: Implications for water quality management in Australian coastal zones, Proceedings of the 27th Hydrology and Water Resources Symposium, Melbourne, May 20-23, 2002.

Presser, T.S., Sylvester, M.A. and Low, W.H. (1994) Bioaccumulation of selenium from natural geologic sources in Western States and its potential consequences, Environmental Management, 18, 423-436.

Queirolo, F., Stegen, S., Restovic, M., Paz, M., Ostapczuk, P., Schwuger, M.J. and Munoz, L. (2000) Total arsenic, lead, and cadmium levels in vegetables cultivated at the Andean villages of northern Chile, The Science of the Total Environment, 255, 75-84.

Schmidt, K.D. (1993) Hydrogeologic factors affecting mobility of trace inorganic constituents, Journal of Irrigation and Drainage Engineering, 119, 600-612.

Smith, J.V.S., Jankowski, J. and Sammut, J. (2001) Vertical distribution of As(III) and As(V) in a coastal sandy aquifer: Marine clays, a potential reservoir for arsenic in coastal groundwater systems, Proceedings of the 10th International Symposium on Water-Rock Interactions, Villasimius, Italy, June 10-15, 2001, 1009-1012.

World Health Organisation (WHO) (1994) Recommendations, second ed., Guidelines for Drinking Water Quality, 1, Geneva.

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SP5o – Chemical Speciation and Modeling

43

The use of solid phase selective extraction techniques to support groundwater chemical data from a coastal aquifer affected by elevated arsenic concentrations

Bethany O’Shea1 and Jerzy Jankowski School of Biological, Earth and Environmental Sciences

University of New South Wales, Sydney NSW, Australia 2052

1PhD Research Student, [email protected]

INTRODUCTIONThe mobility of arsenic in the Stuarts Point coastal aquifer is a critical element in the

determination of the aquifer’s arsenic geochemical cycle. The ease with which arsenic can be released, transported and redistributed through the aquifer has effects on both the anthropogenic use and the environmental receptors of this groundwater resource. Aquifer groundwater chemistry is complex and suggests several geochemical processes may be controlling arsenic distribution.

Sequential extractions have been carried out on sediments collected from the aquifer. The aim of these extractions was to determine sources and/or sinks of arsenic in the aquifer matrix, and their potential to release arsenic in dissolved form, for transport and redistribution throughout the aquifer. A thorough understanding of the solid phase arsenic geochemistry obtained through sequential extraction analyses can then be used to assess groundwater chemical processes contributing to arsenic transport in the aquifer.

STUDY LOCATION AND GROUNDWATER USE The Stuarts Point coastal aquifer is located approximately 400km north of Sydney, Australia.

The aquifer is extensively used for both domestic town water supplies and for the irrigation of local crops. Elevated levels of dissolved arsenic in the groundwater system (>300 g/L) prompted the installation of an arsenic treatment plant. Residents are now provided with drinking water containing <7 g/L of arsenic, the limit set in the Australian Drinking Water Guidelines (NHMRC, 1996).

Groundwater used for the irrigation of local crops (potatoes, avocadoes, stone fruits), however, remains untreated. Some concerns exist over the potential uptake and accumulation of arsenic in these crops. Oyster beds farmed for human consumption in the nearby Macleay estuary are also possibly at risk. Potential discharge of high arsenic groundwater into the estuary may lead to accumulation of this element in the oysters.

This study forms part of a three year project to determine the source, mobilisation, transport and sink(s) of arsenic in this coastal aquifer. The expected benefits of this project include:

a thorough understanding of the arsenic geochemical cycle in the aquifer;assessment of the current management of the aquifer as an important water resource; effects of applying arsenic contaminated irrigation water to local crops; andawareness of potential arsenic impacts on surrounding groundwater dependent ecosystems.

POTENTIAL SOURCES AND/OR SINKS OF ARSENIC IN THE STUARTS POINT COASTAL AQUIFER

There are several possible sources and sinks of arsenic in this coastal aquifer, including, but not necessarily limited to:


SP5o – Chemical Speciation and Modeling

44

Loosely bound to mineral surfaces – Arsenic may be attached to mineral surfaces, and subsequently released due to changes in pH or competition with other oxyanions for surface sorption sites. Oxide minerals and oxyhydroxide grain coatings present in the aquifer are suspected of sorbing labile arsenic. Clay deposits of marine/estuarine origin may also be sorbing arsenic to their outer edges. The abundance of shell material (calcite) in these coastal sediments is another potential contributor to the sorption of arsenic. Co-precipitated within the mineral structure – Iron and manganese oxides may dissolve under aquifer reducing conditions, releasing non-labile arsenic into solution. In contrast, sulphide minerals which are oxidised (such as pyrite contained within acid sulphate soils of this coastal zone) can also release arsenic. Sourced from surrounding mineralised bedrock - Nearby bedrock outcrops (suspected to form part of the basement beneath the aquifer) consist of adamellite granitic intrusions with associated zones of mineralisation, grading from molybdenum to silver-arsenic deposits (Gilligan et al., 1992). The contribution of groundwater flow to the Stuarts Point coastal aquifer from this mineralised bedrock is not well understood, but is considered a potential source of arsenic input to the coastal aquifer. In addition, the aquifer contains dispersed lithic fragments. These lithic fragments have been transported to the coast via ancient river systems originating in the upper reaches of the catchment, and thus are a potential source of arsenic within the aquifer matrix.

SOLID PHASE SEQUENTIAL EXTRACTIONS A sequential extraction technique specific for arsenic has been applied to aquifer sediments.

Fractions were chosen in order to target the potential sources and/or sinks of arsenic identified above: arsenic loosely bound to mineral surfaces; arsenic associated with calcite; arsenic contained within the reducible fraction, ie. iron and manganese oxyhydroxides; arsenic associated with oxidisable materials such as sulphide minerals and organic matter; and the residual mineral fraction.

CONCLUSIONS The results of these sequential extractions will be presented at the conference, with discussion

on the advantages and disadvantages of using these analytical techniques in the determination of the arsenic geochemical cycle. Information obtained from the solid phase arsenic geochemistry will be used to support complex groundwater chemical data. In particular, arsenic mobility in this system is a major concern, and thus was addressed via these analytical techniques, to assess the availability of this toxic element to oysters, crops, humans, and other receptors in the surrounding environment.

REFERENCES Gilligan, L.B., J.W. Brownlow, R.G. and H.F. Henley. 1992. Dorrigo-Coffs Harbour 1:250 000

Metallogenic Map SH/56-10, SH/56-11: Metallogenic Study and Mineral Deposit Data Sheets. New South Wales Geological Survey, Sydney.

National Health and Medical Research Council (NHMRC) & Agriculture and Resources Management Council of Australia and New Zealand. 1996. Australian Drinking Water Guidelines. National Water Quality Management Strategy, Canberra.

UNCOVERING THE SOURCE AND DISTRIBUTION OF ELEVATED ARSENIC IN A COASTAL AQUIFER: CLUES FROM THE AQUIFER MATRIX

Bethany O’Shea: School of Biological, Earth & Environmental Sciences, University of New South Wales, Sydney NSW 2052 Australia.

Jerzy Jankowski: School of Biological, Earth & Environmental Sciences, University of New South Wales, Sydney NSW 2052 Australia.

Introduction

Elevated arsenic in groundwater is a common occurrence throughout many parts of the world, including Argentina, Bangladesh, Chile, Hungary and many parts of the USA (Smedley et al., 2002; Welch et al., 1999). Arsenic concentrations in aqueous form can be derived from a number of sinks and sources, both natural and anthropogenic. The weathering of arsenic containing minerals; occurrence of geothermal waters; oxidation of acid sulfate soils incorporating arsenic in their structure; and reductive dissolution or desorption processes releasing arsenic from the solid phase in the aquifer matrix, can naturally elevate arsenic concentrations in groundwater. Anthropogenic arsenic contamination can include pesticide application, mining activities, sheep and cattle dipping chemicals and a variety of industrial processes.

Some of the most common theories proposed for the source and mobilisation of naturally elevated arsenic concentrations in groundwater are:

the oxidation of sulfides and associated release of arsenic into solution; the reductive dissolution of oxides (and release of sorbed arsenic) driven by natural organic matter present in the aquifer; pH influenced desorption from iron/manganese/aluminium oxides; desorption of arsenic from other mineral surfaces, such as clays; and competitive adsorption of phosphate and other anions for the same arsenic sorption sites present within the aquifer matrix.

Many studies have identified arsenic in the aqueous phase and postulated which of the above geochemical processes is responsible for controlling its distribution - based on chemical trends present in the water data. It is often not possible to study the interaction between the aquifer matrix and the aqueous phase due to lack of sediment data or the cost associated with drilling new wells and collecting and analysing the sediments in conjunction with groundwater. This study was designed to use both groundwater and sediment data to complement each other and provide information on the distribution of arsenic between each phase. A selection of solid phase analyses are discussed herein to indicate possible arsenic sinks and sources present within the aquifer matrix.

Fig. 1. Stuarts Point aquifer.

Study Location and Geology

Stuarts Point is located on the New South Wales Mid-North coast (Fig. 1). Groundwater for town and farm use is extracted from a Holocene coastal sand aquifer consisting of intermittent layers of estuarine sands and clays. Bedrock beneath these unconsolidated deposits is inferred to consist of marine

Inaugural Australasian Hydrogeology Research Conference 2004 87

Permian rocks of the Kempsey Beds and the Lower Palaeozoic Nambucca Beds, separated by the east-west trending Kempsey Fault. The Kempsey Beds consist of sandstones, slates, mudstones and tuffs (Voisey, 1958) and have been downthrusted by the Kempsey Fault, relative to the older Lower-Palaeozoic Nambucca Beds. These latter beds consist of slates and phyllites, which have been intruded by coastal granitoids to form nearby Mt Yarrahapinni. Considerable mineralisation occurs within Mt Yarrahapinni, grading from an outward zoning molybdenum zone to a silver arsenic zone (Gilligan et al., 1992). Stibnite deposits are also common in the region.

Arsenic in the Groundwater

The surrounding mineralisation within the regional geology is an obvious source of arsenic if a geological/hydrological link can be identified between known mineral occurrences and the aquifer. Alternatively, Acharyya et al. (1999) proposed that arsenic became preferentially entrapped in fine-grained and organic-rich sediments during mid-Holocene sea-level rises in the Bengal Basin; a theory which could also apply at Stuarts Point. Anthropogenic influences cannot be ruled out entirely since pesticide application and cattle/sheep dipping may have occurred in the area.

Arsenic concentrations in Stuarts Point groundwater have been reported up to 47 times greater than the 7 g/L limit for As in Australian drinking water (Smith et al., 2003). These previous investigations and those carried out by the authors herein have deduced that no single geochemical process is responsible for arsenic release into the aqueous environment. A good correlation can exist between As and pH if the dominant control over As is pH influenced desorption from mineral surfaces. Fig. 2a shows arsenic appears to be independent of pH and as such the dominant control must not be pH influenced sorption processes. Fig. 2b delineates several sample points that show good correlation between As and SO4

2-

suggesting these samples may be influenced by pyrite oxidation and subsequent release of As; or seawater intrusion into the freshwater aquifer. Fig.’s 2c and 2d show very minor (if any) correlation between Fe and As and HCO3 and As, dismissing dominance of organic matter driven iron oxide reduction as a mechanism of As release. This data shows that there is no single dominant process controlling As mobility in the Stuarts Point aquifer.

Fig. 2. (a) A good correlation does not exist between As and pH. Arsenic may be mobile in slightly acidic and neutral-slightly basic waters. (b) Some waters show a positive correlation between As and SO4

2- indicating pyrite oxidation or seawater intrusion may influence arsenic distribution. (c) No correlation between Fe and As or (d) between HCO3

- and As, dismissing the dominance of oxide reduction and subsequent As release.


Arsenic in Solid Phase

Given the uncertainty regarding controls on As distribution suggested from water data alone, it was necessary to analyse the solid phase of the aquifer matrix to determine sources and sinks of As. Sequential extraction procedures, XRF, XRD, statistical analysis and SEM techniques were used to ascertain whether or not postulated sinks were present in the aquifer matrix. Targeted sinks included:

the presence and identification of specific clay types in the matrix and their potential to adsorb As onto their outer edges; any specific As mineral phases such as scorodite, arsenopyrite or barium arsenate; correlations between As and other elements in the solid phase; influence of shell material on the sequestration of As; the presence of pyrite and its ability to incorporate As into its crystal structure; and the presence of iron/aluminium/manganese oxides and oxyhydroxides and their ability to sorb As.

Analysis of the solid phase enabled correlations to be made between various trace elements present in the matrix, leading to a hypothesised source of arsenic for the aquifer. Further high powered microscopic analysis identified potential sinks of arsenic redistributed throughout the matrix. Water chemistry was used to ascertain which chemical processes are currently active in the aquifer and therefore which processes are controlling As mobility. Due to the large amount of graphics generated during this analysis process the results will not be incorporated within this abstract, rather they will be presented at the conference where further discussion can warrant their importance.

Conclusion

Arsenic has been identified at elevated concentrations in a coastal aquifer. Investigations were initiated to determine whether the As was naturally occurring or anthropogenically induced. Results from groundwater sampling alone were inconclusive and suggested many geochemical processes were controlling As distribution in the aquifer. A thorough analysis of the solid phase of the aquifer matrix has provided a possible source for the As in the aquifer, in addition to identifying current geochemical sinks present within the aquifer. This integrated approach to the investigation has allowed a better understanding of As geochemical processes to be determined and stresses the importance of considering the solid phase component of any groundwater study.

References

Acharyya SK, Lahiri S, Raymahashay BC, Bhowmik A (1999) Arsenic toxicity of groundwater in parts of the Bengal basin in India and Bangladesh: the role of Quaternary stratigraphy and Holocene sea-level fluctuation. Environmental Geology 39:1127-1137.

Gilligan LB, Brownlow JW, Cameron RG, Henley HF (1992) Dorrigo-Coffs Harbour 1:250 000 Metallogenic Map SH/56-10, SH/56-11: Metallogenic Study and Mineral Deposit Data Sheets. New South Wales Geological Survey, Sydney.

Smedley PL, Nicolli HB, Macdonald DMJ, Barros AJ, Tullio JO (2002) Hydrogeochemistry of arsenic and other inorganic constituents in groundwaters from La Pampa, Argentina. Appl. Geochem. 17:259-284.

Smith JVS, Jankowski J, Sammut J (2003) Vertical distribution of As(III) and As(V) in a coastal sandy aquifer: factors controlling the concentration and speciation of arsenic in the Stuarts Point groundwater system, Northern New South Wales, Australia. Appl. Geochem. 18:1479-1496.

Voisey AH (1958) Tectonic evolution of North-Eastern New South Wales, Australia. Journal of the Royal Society of New South Wales, 92: 191-203.

Welch AH, Helsel DR, Focazio MJ, Watkins SA (1999) Arsenic in ground water supplies of the United States. In: Chappell WR, Abernathy CO, Calderon RL (Eds.), Arsenic Exposure and Health Effects. Elsevier, Amsterdam, 9-17.


8th ICOBTE, Adelaide, Australia, April 4-9 2005

Arsenic Cycling in a Coastal Aquifer: Uncovering the Source, Sinks and Mobilisation Processes

Bethany O’Shea1 and Jerzy Jankowski1UNSW Groundwater Group, School of Biological, Earth and Environmental Sciences,

University of New South Wales, Sydney, NSW 2052, AUSTRALIA ([email protected] )

INTRODUCTIONArsenic concentrations above Australian Drinking Water Guidelines (7 g/L) have been

reported in various aquifers along the New South Wales (NSW) coast. Groundwater investigations at Stuarts Point, on the Mid-North coast of NSW, have reported arsenic concentrations in excess of 330 g/L (Smith et al., 2003). This study was initiated to determine if the arsenic occurrence is natural or anthropogenic, and which geochemical processes are governing its distribution. Initial hypotheses linked the source of arsenic to Acid Sulfate Soils (ASS) and Holocene marine clays. Pyrite, one of the main constituents of ASS, can incorporate arsenic into its crystal structure and hence release it upon oxidation. Illitic clays, deposited in the aquifer during Holocene sea level fluctuations, have the ability to sorb arsenic onto their outer edges and subsequently release it when geochemical conditions are conducive. Results from current investigations now indicate these hypothesised geochemical processes may be utilised as current arsenic sinks within the aquifer, but are not responsible for its initial presence in the aquifer (ie, arsenic source). An additional aim of this study is to determine whether or not elevated arsenic is likely to be present throughout other NSW coastal environments with similar geomorphic and depositional histories; or if the occurrence is simply a localised problem.

METHODSXRD, XRF, SEM, sequential extractions and statistical analysis of the sediments were

conducted to ascertain whether a plausible source of the arsenic could be determined, and investigate potential arsenic sinks in the solid phase. Groundwater data enabled an assessment of geochemical conditions suitable for arsenic partitioning between solid and aqueous phase. Arsenic cycling through irrigation water, accumulation in farm soils and vegetables, and discharge to the adjacent estuarine environment were each investigated to determine the impacts of this toxic element on humans and the receiving ecosystem.

RESULTS AND DISCUSSION Initial investigations of groundwater suggested arsenic concentrations were being

influenced by a series of complex hydrogeochemical processes. Aqueous data presented in Figure 1 supports this conclusion; no obvious correlation exists between arsenic and known geochemical processes thereby excluding the dominance of one singular controlling release mechanism.

This irregular distribution of aqueous arsenic concentrations warranted investigations into the source of arsenic in the solid phase of the aquifer matrix. XRD did not identify any significant arsenic mineral phases and XRF reported arsenic concentrations in solid phase no higher than 14ppm. Statistical correlations were observed between arsenic and antimony in the sediments, in addition to arsenic and molybdenum in the groundwater. Both of these elements are present within regional mineralised geology. A thorough geomorphic reconstruction and analysis of the relationship between indicator trace elements and arsenic


8th ICOBTE, Adelaide, Australia, April 4-9 2005

suggest the original provenance of arsenic in the aquifer is likely to be derived from local and/or regional stratigraphy. Regional geologic influences are hypothesised to be associated with fluvially transported sediments now forming parts of the aquifer matrix. For example, stibnite deposits located in the upper catchment of the Macleay River have oxidised and contributed dissolved antimony and arsenic to the Macleay River (Ashley et al., 2003), illustrating the past and present potential for contribution of trace elements to the aquifer matrix via river processes. Alternatively, an in-situ upwards vertical gradient may be contributing arsenic and trace elements from the underlying mineralised bedrock to the unconsolidated aquifer above.

Fig. 1. (a) A good correlation does not exist between As and pH, indicating pH influenced desorption processes are not dominant. Arsenic may be mobile in slightly acidic and neutral-slightly basic waters. (b) Some waters show a positive correlation between As and SO4

2- indicating pyrite oxidation or seawater intrusion may influence arsenic distribution in these samples. (c) No correlation between Fe and As dismissing the dominance of an oxide reduction and subsequent As release mechanism.

While the source of arsenic is still being hypothesised, current sinks within the aquifer matrix have successfully been identified. Arsenic-rich pyrite, iron oxyhydroxides and illitic clay coatings on aquifer particles suggest several arsenic sinks exist in the aquifer, each contributing to arsenic occurrence and its variable distribution. This explains the poor relationships observed in the aqueous data presented in Figure 1. Investigations are currently under way to determine the aqueous conditions conducive to arsenic mobilisation within the groundwater and whether anthropogenic activities are enhancing its release from solid phase.

CONCLUSIONS Arsenic is therefore suggested to be naturally derived from local and/or regional geology.

Its mobility in this drinking water aquifer is controlled by several geochemical processes, as identified by SEM and aqueous chemical interpretation methods. Arsenic cycling through irrigation water does not cause accumulation of this element in soils or vegetables, however potential arsenic discharge to the adjacent estuary is of concern to local oyster production. The source of arsenic in this coastal environment is thought to be localised according to geology, however, given its ubiquitous nature, arsenic should not be ignored in other NSW coastal environments and therefore should be investigated accordingly.

REFERENCES Smith, J.V.S., Jankowski, J. and Sammut, J. (2003) Vertical distribution of As(III) and As(V)

in a coastal sandy aquifer: factors controlling the concentration and speciation of arsenic in the Stuarts Point groundwater system, Northern New South Wales, Australia. Appl.Geochem. 18:1479-1496.

Ashley, P.M., Craw, D., Graham, B.P. and Chappell, D.A. (2003) Environmental mobility of antimony around mesothermal stibnite deposits, New South Wales, Australia and southern New Zealand. J. Geochem. Expl. 77:1-14.

NATURALLY ELEVATED ARSENIC AND OTHER TRACE ELEMENTS IN A SAND AQUIFER OF EASTERN AUSTRALIA

O’SHEA, Bethany. Department of Geology, Dickinson College, PO Box 1773, Carlisle, PA 17013, [email protected] and JANKOWSKI, Jerzy. School of Biological, Earth & Environmental Sciences, University of New South Wales, Sydney 2052, Australia.

Aquifers comprised of unconsolidated sediments can be valuable resources due to their high groundwater yields. Much of eastern coastal Australia contains sand aquifers that are heavily exploited for drinking water and irrigation. Unless specifically targeted, many trace elements and metals have not been part of routine analyte suites in groundwater risk assessments in past decades. The recent widespread occurrence of arsenic in many sedimentary aquifers of the world prompted the analysis of dissolved arsenic in a drinking water supply aquifer at Stuarts Point, 600 km north of Sydney. Arsenic has been reported above the Australian drinking water limit of 7 g L-1 in 49% of groundwater samples analyzed. A detailed hydrochemical characterization of the aquifer revealed the source of arsenic to be natural, weathered from a mineralized arsenic-bearing stibnite (Sb2S3) deposit in the hinterland and transported downstream by fluvial processes.

Also reported above drinking water guidelines in this seemingly ‘pristine’ and low contaminant risk environment were Cd, Hg, Mn, Pb and Se. However, spatial distribution is not widespread like dissolved arsenic distribution. Multivariate statistical analyses have linked these elevated elemental occurrences to specific zones in the aquifer. Such zones include shallow groundwater subjected to pyrite oxidation, groundwater flowing through fluvially deposited sediments or mineralized weathered bedrock clays, and groundwater impacted by seawater intrusion processes. Variation in chemical heterogeneity, particularly redox conditions and groundwater-matrix interaction, govern the sporadic distribution of these elevated elements.

These results can aid in determining optimal depths of groundwater withdrawal for various uses in the Stuarts Point aquifer. For example, Cd occurs in elevated concentrations at approximately 10 m below the ground surface. Bores installed primarily for irrigation of potato crops should therefore target shallow groundwater (less than 10 m) in order to reduce the possibility for Cd accumulation in potatoes. Distinct groundwater zones occur within a very limited depth (30 meters) in this aquifer, illustrating the importance of considering chemical heterogeneity in groundwater risk assessments for all aquifer matrices.


A comparison of arsenic occurrence and geochemistry in two coastal

groundwater environments B.M. O’SHEA1, G. CLARK1 AND J. JANKOWSKI2

1 Department of Geology, Dickinson College, Pennsylvania, USA; [email protected]; [email protected]

2 School of Biological, Earth & Environmental Science, University of New South Wales, Australia; [email protected]

Arsenic occurrence and geochemistry in two distinctly different coastal aquifers has been investigated. The first, an unconsolidated sand and clay aquifer of eastern Australia, has been compared to groundwater composition contained within sand and gravel surficial deposits and bedrock of coastal Maine, USA. Both studies investigate the presence of natural arsenic in groundwater being used for potable supply. The Australian aquifer, located in northern New South Wales, has been investigated in detail [1]. Dissolved arsenic concentrations are greater than 300 g L-1. Arsenic bearing stibnite deposits in the upper catchment have been fluvially eroded and transported over time to deposit an aquifer lithology elevated in arsenic. Current groundwater conditions are suitable for both arsenic mobilisation and immobilisation processes, producing a cyclic control on arsenic distribution. The redox state of the groundwater largely controls arsenic distribution. Reductive dissolution of iron oxyhydroxides releases arsenic into shallow parts of the aquifer (0-20m depth). As redox conditions become more anoxic with depth, the reductive dissolution of iron oxyhydroxides approaches completion and correlates with increased arsenic concentrations in groundwater. At depths of 25-30m, seawater intrusion draws increased sulfate concentrations into the fresh water aquifer and promotes reduction of sulfate and the subsequent precipitation of iron sulfides. Geochemical modeling and redox conditions support this hypothesis. Arsenic concentrations in groundwater signficantly decrease as iron sulfide minerals incorporate the arsenic into their mineral structure. Arsenian pyrite has been identified with an Electron Microprobe. In comparison, the sporadic occurrence of arsenic in coastal Maine is proposed to derive from the oxidation of arsenical pyrite in fractured igneous and metamorphic bedrock of mid-coastal Maine [2]. Investigations continue in an extended area of coastal Maine to complement the work conducted by [2] and allow a detailed geochemical comparison to be conducted on two distinctly different coastal aquifers bearing similar aqueous geochemical signatures and associated health risk problems. References [1] O’Shea B.M. (2006) PhD Thesis. University of NSW. Australia. [2] Sidle W.C., Wotten B., Murphy E. (2001) Env.Geol. 41, 62-73.




Geogenic arsenic in an Australian sedimentary aquifer: risk awareness for aquifers in Latin American countries

B. O’Shea Department of Geology, Dickinson College, Carlisle PA, USA

J. Jankowski School of Biological, Earth and Environmental Sciences, The University of New South Wales, Sydney NSW, Australia

ABSTRACT: Arsenic concentrations in a coastal sandy aquifer used extensively for drinking water and irrigation supply have been reported at more than 45 times the permissable limit in Australian drinking water. The presence of arsenic in the Stuarts Point aquifer was surprising due to its assumed deposition by on-shore sediment supply and thus lack of an obvious arsenic source. Development of an aquifer-specific geomorphic model now suggests that arsenic has been derived from regional erosion of arsenic-rich stibnite deposits. The heterogeneity of the sedimentary aquifer causes groundwater redox conditions to control arsenic mobilisation via re-ductive dissolution of iron oxyhydroxides and precipitation of arsenian pyrite. The valuable les-sons learned from an Australian sedimentary aquifer where arsenic occurrence is complex, should be utilised in the assessment and identification of potential arsenic-rich groundwaters of Latin America.

1 INTRODUCTION

In the last decade there has been a dramatic increase in the number of studies investigating arse-nic in groundwater. Much of this can be attributed to the discovery of elevated arsenic in domes-tic groundwater supplies in Bangladesh (Nickson et al., 2000; Harvey et al., 2005). The in-creased awareness of the ubiquitous nature of arsenic in the environment led other countries to focus on the occurrence of arsenic in their groundwater. Elevated arsenic has since been found in groundwater of Australia (Smith et al., 2003; 2006, O’Shea et al., 2006), Taiwan (Chen et al., 1994), Vietnam (Berg et al., 2003) and many parts of the USA (Schreiber et al., 2000; Welch et al., 2000; Sidle et al., 2001). In recent years, the occurrence of arsenic in aquifers of Latin American countries has become increasingly apparent (Smedley et al., 2002; Bundschuh et al., 2004; Bhattacharya et al., 2006; Romero et al., 2004).

The effects of ingesting arsenic-rich groundwater can be devastating both physically and so-cially. Physical effects include various cancers (Bissen and Frimel, 2003), skin lesions, gangrene and other clinical symptoms like joint pain, chronic cough and abdominal pain (Khalequzzaman et al., 2005). Unfortunately, many people who depend on arsenic-rich groundwater supplies also live in rural areas where basic necessities are unsatisfied and incidences of poverty can be high, such as the Chaco-Pampean Plain of Argentina (Bundschuh et al., 2004).

In contrast, the Stuarts Point aquifer of eastern Australia supports small coastal communities where the standard of living is more than satisfactory. Dissolved arsenic has been reported by Smith et al. (2003) at concentrations as high as 337 g/L, more than 45 times the acceptable drinking water limit of 7 g/L As (NHMRC, 1996). As such, groundwater in this aquifer is now being treated to remove dissolved arsenic prior to human consumption.

The purpose of this study is to provide an outline of the investigation conducted at Stuarts Point. Arsenic sources, sinks and controls on mobilization into groundwater have been exam-ined in detail (O’Shea, 2006) and are described herein. By examining the geochemical controls

on arsenic occurrence in this aquifer, it is hoped that the outcomes produced in this study can be applied to the assessment of aquifers at risk of arsenic occurrence in Latin American countries.

2 ENVIRONMENTAL SETTING 2.1 Aquifer Lithology and Hydrogeology Stuarts Point is located approximately 400 km north of Sydney, on the New South Wales (NSW) mid-north coast (Figure 1). Groundwater for domestic supply and agricultural irrigation is extracted from the Stuarts Point aquifer.

Figure 1. Location of the Stuarts Point aquifer and multi-level piezometers installed for this study. The-borelog from ML9 shows the general stratigraphy of the aquifer sediments.

The aquifer is approximately 20-50 m thick and consists of unconsolidated sediment overly-ing regional bedrock geology. Previous investigations (Eddie, 2000; Smith et al., 2003) have predominantly described the aquifer as a beach barrier deposit, with little emphasis placed on fluvial contribution. It is now understood (O’Shea, 2006) that a combination of fluvial, estuarine and onshore sediment supply have contributed to aquifer formation during the Quaternary Pe-riod. This combination of depositional conditions accounts for the heterogenous nature of the aquifer lithology, as seen in the generalized stratigraphy shown in Figure 1. Four distinct lithologic facies, “beach barrier sand”, “fluvial sand”, “fluvial sand / estuarine clay” and “bed-rock clay” can be identified throughout the aquifer.

Groundwater in the unconsolidated aquifer flows in a south-easterly direction from the topog-raphic high at the base of Mt Yarrahapinni towards the Macleay Arm estuary located adjacent to the aquifer. Hydraulic conductivity varies through the heterogenous sand and clay units, ranging from 0.01 to 35.3 m/day (Northey, 2001). The aquifer is recharged directly via precipitation and indirectly via leakage from the bedrock aquifer below.

2.2 Regional and Bedrock Geology Regional and bedrock geology belongs to the New England Fold Belt (NEFB). Part of the NEFB is drained by the Macleay River, which discharges to the ocean immediately south of the Stuarts Point aquifer. Nearby Mt Yarrahapinni (Figure 1) is a coastal granitoid with mineraliza-tion grading from an outward zoning molybdenum zone, to a silver-lead zone and finally to a silver-arsenic zone (Gilligan et al., 1992). It is possible that these zones may extend beneath the northern portion of the Stuarts Point aquifer.

3 METHODOLOGY3.1 Sample Collection and Preservation Ten boreholes were drilled by solid flight augers and later converted to multi-level piezometers for groundwater sampling. During drilling (July 2001), 36 sediment samples were collected at 1.5 m intervals via in-situ split spoon sampling in four multi-levels (ML7 – ML10). Sediment samples were field-logged, frozen and kept in the dark during transportation to the laboratory. Groundwater in the ten multi-level piezometers (ranging in depth from 22 m to 30 m with a sample point located at 1-metre depth intervals) was purged using a peristaltic pump. A total of 227 samples were collected when general parameters (pH, temperature, electrical conductivity) had stabilised to within +/- 5% (O’Shea, 2006). Samples were filtered through 0.45 m Milli-poreTM cellulose acetate membrane filter paper, preserved with concentrated analytical grade ni-tric acid and kept chilled until analysis in the laboratory.

3.2 Sediment and Groundwater Analysis X-Ray Fluorescence (Siemens SRS300) was conducted at the University of New South Wales to determine total elemental composition of the sediments. Scanning Electron Microscopy (SEM) was performed at the Electron Microscope Unit at the University of New South Wales on a Hi-tachi S4500 Field Emission SEM (1996), with high resolution (1.5 nanometers), a tilting stage, Robinson Back-Scatter Detector, Oxford Cathodoluminescence Detector (MonoCL2/ISIS) and a Link ISIS 200 Microanalysis System for chemical determination. The accelerating voltage was set at 20kV throughout the analysis process. Arsenic was not expected to be detected during SEM analysis given the highest reported As concentration (14 mg kg-1) by XRF was well below the As detection limit (approximately 50 mg kg-1) for the SEM. A Cameca SX50 electron mi-croprobe engaging four Wavelength Dispersive Spectra’s and one Energy Dispersive Spectra, giving it the advantage of a more accurate and sensitive chemical analysis than the standard SEM-EDS analysis, was also employed. Trace elements in groundwater were analysed by ICP-MS and major ions were analysed by ICP-AES in the chemical laboratory at the University of New South Wales. Bicarbonate was determined in the field by titration with 0.01 M HCl against bromocresol green indicator (American Public Health Association, 1992). Statistical analyses (descriptive statistics and Pearsons correlations) were performed in SPSS version 12.01 statisti-cal software package as described in O’Shea and Jankowski (2006).

3.3 Quality Assurance/Quality Control Duplicate sediment and groundwater samples were collected during the field program. The ac-curacy of the sediment analyses was assessed by running duplicate XRF analyses. Duplicate re-sults were reported within +/-1%. For arsenic, a Japanese Reference Material (JB-1) was used to ensure solid arsenic concentrations were reported within +/-1%. Both microprobe instruments were calibrated with Standard Reference Materials for the various elements analysed. The S4500 uses Guide E1508-98 Standard Guide for Quantitative Analysis by Energy-Dispersive Spectroscopy; while the SX50 employs the data reduction matrix correction procedure based on the methods of Pouchou and Pichoir (1985). Charge balance error (CBE) values for the ground-water samples were generally in the range of +/-2-3%. CBE’s above +/-12% were rejected. The interference between argon and chloride during the analysis of arsenic by ICP-MS can raise re-ported arsenic concentrations by 1 g/L for every 100 mg/L of chloride present. Chloride con-

centrations in groundwater sampled for this study reported an average concentration of 42 mg/L. Interference from the formation of argon chloride should thus be minimal in these samples. Field titrations for bicarbonate were repeated at the well-head and the average of the two results taken.

4 ARSENIC OCCURRENCE AND DISTRIBUTION AT STUARTS POINT

4.1 Source of Arsenic in the Aquifer The presence of arsenic in the Stuarts Point aquifer was surprising due to its assumed deposition by on-shore sediment supply and thus lack of an obvious arsenic source. Coastal aquifers along the east coast of Australia have endured similar depositional conditions and eustatic changes in sea level, initially raising concern over the likely presence of arsenic in other coastal aquifers (O’Shea and Jankowski, 2002; Smith et al., 2006). In the absence of any significant anthropo-genic sources the arsenic at Stuarts Point is deemed naturally occurring. Four geogenic sources are proposed:

Arsenic has been contributed to the aquifer matrix via deposition of regionally eroded geo-logical units containing arsenic mineralization (O’Shea et al., 2006); Arsenic is derived from remnant seawater trapped in marine clay units deposited during eustatic changes of sea level in the Quaternary (Smith et al., 2006); The oxidation of arsenian pyrite present in ASS material contributes dissolved arsenic to the groundwater (O’Shea, 2006; Smith et al., 2006); and/or The underlying bedrock contains arsenic, which is being contributed to the aquifer via up-wards vertical leakage of groundwater (O’Shea, 2006; Smith et al., 2003).

The dominant arsenic source supported herein is derivation of arsenic from regional erosion of arsenic-rich stibnite deposits in the upper reaches of the Macleay River (O’Shea et al., 2006). A combination of sediment chemistry, statistical analysis, palaeontological interpretation and sedimentological analysis allowed the construction of a detailed geomorphic model specific to the Stuarts Point aquifer (Figure 2).

This model confirms fluvial sedimentation occurred in the aquifer and thus prompted the in-vestigation into an upgradient arsenic source. Naturally eroded arsenic-rich stibnite (Sb2S3) de-posits from the upper catchment have been linked to the arsenic currently present in the Stuarts Point aquifer sediments (O’Shea et al., 2006). Minor contributions of arsenic from other sources can be expected in localized areas of the aquifer, particularly as discharge from mineralized bed-rock in the northern portion of the aquifer (O’Shea et al., 2006).

4.2 Current Arsenic Sinks Mean sediment element concentrations are listed in Table 1. Arsenic concentrations in the aqui-fer matrix range from 1.4 to 14.0 mg kg-1 and represent average background values. Quartz grains dominate the sand aquifer and are responsible for high concentrations of silica in the ma-trix. Aluminium and iron concentrations increase in clay facies.

Frequently associated with naturally occurring arsenic are aluminium, silica and potassium (clay minerals); iron and manganese (oxides); and sulfur (commonly delineating a pyrite asso-ciation). Thus, a correlation between arsenic and one or more of these elements could identify the presence of particular solid phase arsenic sinks. Pearsons correlations (Table 2) show statis-tical associations between arsenic and these elements in each lithologic facies.

Figure 2. Geomorphic model proposed for Stuarts Point by O’Shea et al. (2006).

Table 1. Selected mean element concentrations for sediments in each lithologi_____________________________________________________________________________ c facies at Stuarts PointElement Barrier Fluvial Fluvial sand/ Bedrock % oxide sand sand estuarine clay clay _____________________________________________________________________________SiO2 96.11 93.06 84.89 72.34 Al2O3 0.94 1.53 5.24 12.37 Fe2O3 0.28 0.54 1.40 4.84 MnO 0.01 0.01 0.02 0.02 MgO 0.21 0.26 0.49 0.87 CaO 0.05 1.59 2.49 0.78 Na2O 0.37 0.46 0.79 0.92 K2O 0.25 0.47 1.13 1.78 SO3 0.04 0.19 0.37 0.59 As* 3.99 6.33 6.99 9.33 _____________________________________________________________________________*Reported in mg kg-1.

Table 2. Correlation (P<0.05) between sediment arsenic concentrations and major elements for each lithologic facies in the Stuarts Point aquifer _____________________________________________________________________________

Correlation with arsenic Element* Barrier Fluvial Fluvial sand/ Bedrock sand sand estuarine clay clay _____________________________________________________________________________SiO2 -0.24 -0.73 -0.57 0.95 Al2O3 0.74 0.65 0.73 -0.98 Fe2O3 0.70 0.93 0.01 0.68 MnO -0.12 0.01 0.16 -0.06 MgO -0.29 0.84 0.14 -0.66 CaO -0.15 0.53 0.26 -0.09 Na2O 0.25 0.76 0.22 -0.57 K2O 0.42 0.70 0.14 -0.98 SO3 0.08 0.78 0.38 -0.09 _____________________________________________________________________________*Analysed by XRF and reported as % oxide.

There is a strong negative correlation with As and the clay mineral indicators, aluminium (r2

= -0.98) and potassium (r2 = -0.98), and a strong positive correlation with silica (r2 = 0.95). Illite group clay minerals contain Al, K and Si and have been identified as coatings on quartz grains (Figure 3), however no arsenic has been identified associated specifically with illite occurrence in the aquifer matrix. Illite has a moderate surface area available for sorption when compared to other clay minerals such as kaolinite. Positive surface charges for anion adsorption can be gen-erated by protonation on broken Al-OH bonds exposed at the particle surface. Lin and Puls (2000) found illite had moderate arsenic adsorption and subsequently moderate desorption. Therefore, arsenic adsorption to clay minerals may be occurring, but is not considered a domi-nant sink within the Stuarts Point aquifer matrix.

Figure 3. Electron Microprobe photograph and quantitative analysis of an illite coating on a quartz grain found within the fluvial sand / estuarine clay lithologic facies. XRD has also identified illite within the matrix (O’Shea, 2006).

Arsenic and iron correlate strongly (r2 = 0.93) in the fluvial sands. These sands are orange-brown in colour, a characteristic which is frequently indicative of iron oxyhydroxides presence. Smedley and Kinniburgh (2002) note that iron oxides are probably the most important adsorb-ents in sandy aquifers because of their great abundance and strong binding affinity. SEM identi-fied iron oxyhydroxide coatings on sand grains in the matrix (Figure 4). It is interesting to note that arsenic and iron are not at all correlated (r2 = 0.01) for the fluvial sand / estuarine clays. These sands and clays are characterized by an olive/grey colour indicating iron oxide minerals are no longer dominant. There is no correlation (r2 < 0.16) between arsenic and manganese in the aquifer, suggesting arsenic is not associated with manganese oxides at Stuarts Point.

Figure 4. SEM photograph and semi-quantitative analysis (within the spectral box) showing Fe oxyhy-droxide presence within the fluvial sands of the Stuarts Point aquifer.

A moderate correlation (r2 = 0.78) exists between arsenic and sulfur in the fluvial sands, sug-gesting the possibility of an arsenic sulfide relationship. Arsenian pyrite was identified by elec-tron microprobe (Figure 5) as discrete micron sized pyrite clusters on quartz grains and shell fragments. These arsenic-rich pyrites are thought to be precipitating under changes in redox conditions of the groundwater (Section 4.3). The numerous statistical correlations observed be-tween arsenic and other elements suggest several geochemical processes are acting on its distri-bution in the solid phase.

Figure 5. Electron microprobe photograph and quantitative analysis shows a discrete arsenian-pyrite clus-ter on a shell fragment in the Stuarts Point aquifer.

4.3 Controls on Arsenic MobilizationThe successful identification of iron oxyhydroxides and arsenian pyrite phases in the aquifer matrix directs groundwater interpretation towards investigating the validity of these two sinks as arsenic mobilization processes within the aquifer.

Groundwater flowing through the orange-red, iron rich, fluvial sands show a moderate (r2=0.54) correlation between dissolved Fe and HCO3 (Figure 6). A good correlation can be ex-pected (1) if iron oxyhydroxides are dissolving under reducing conditions (-50 to -150 mV in this part of the aquifer).

8FeOOH(-As) + CH3COOH + 14H2CO3 8Fe2+ + 16HCO3- + As + 12H2O (1)

The abundant shell material dissolving in the Stuarts Point aquifer may influence this correla-tion as excessive amounts of HCO3 are added during calcite dissolution. Arsenic is expected to be released from dissolving iron oxyhydroxide minerals but re-adsorbed onto fresh surfaces, thus keeping dissolved arsenic concentrations low (mean As for these groundwaters is 10.9

g/L). This correlation decreases in the underlying olive grey fluvial sand/estuarine clays (r2=0.44) where iron oxyhydroxide dissolution approaches completion and arsenic concentra-

tions begin to increase (maximum 61.6 g/L As) as adsorption sites in the aquifer matrix de-crease.

Figure 6. Total Fe versus HCO3- in groundwater flowing through the fluvial sands of the aquifer matrix.

The formation of the identified arsenian pyrite in the aquifer is also a possible control on arsenic mobilization in dissolved phase. Aqueous redox potential is sufficiently reducing (up to -235 mV in some deeper parts of the aquifer) to promote SO4

2- reduction to S2- under natural condi-tions. Marine estuarine clays may be the source of organic matter and sulfate; with addition of iron and arsenic from the dissolution of iron oxides (2).

4SO42- + Fe2O3 + 8CH2O 2Fe(As)S2 + 8HCO3

- + 4H2O (2)

Once palaeowater salinity has exhausted the supply of sulfate, pyrite precipitation may be lim-ited. This process is bacterially mediated and is the subject of much current research (Oremland and Stolz, 2005). Localized seawater intrusion contributes additional sulfate to the fresh water aquifer and increases arsenian pyrite formation.

Several other processes have been found to effect arsenic mobilization in the Stuarts Point aquifer and are currently under further investigation. These include arsenian pyrite oxidation by anthropogenic nitrate and addition of arsenic from the adjacent Macleay Arm estuary under seawater intrusion processes (O’Shea, 2006). The overall mobility of arsenic in the aquifer is in-fluenced by both anthropogenic (moderate contribution) and natural processes (dominant contri-bution) and is largely controlled by redox conditions and aquifer heterogeneity.

5 IMPLICATIONS FOR LATIN AMERICAN AQUIFERS

Emerging research suggests multiple geogenic sources may contribute to arsenic occurrence in Latin American aquifers. Bundschuh et al. (2004) suggested three sources of arsenic to ground-water in the Chaco-Pampean Plain of Argentina:

Layers of volcanic ash with 90% rhyolitic glass; Volcanic glass dispersed throughout the sediments; and Clastic sediments of metamorphic and igneous origin.

They concluded that the clastic sediments, which are the source of arsenic at Stuarts Point, con-tribute little or no As to groundwater in the Chaco-Pampean Plain. Also in contrast to the inves-tigations described herein are the redox geochemical conditions observed in Latin American aq-uifers. Arsenic has been found in oxidizing groundwater environments (Del Razo et al., 1990; Rodriguez et al., 2004) rather than reducing environments like Stuarts Point and Bangladesh.

What is similar between most arsenic groundwater occurrences are the dominant mobiliza-tion processes releasing arsenic to the environment. At Stuarts Point, aquifer heterogeneity causes several of these release mechanisms to occur; oxidation of pyrite (O’Shea, 2006), reduc-tive dissolution of iron oxyhydroxides, pH-influenced desorption (Smith et al., 2006) and pre-cipitation of arsenian pyrite. While it is important to delineate high arsenic groundwater re-sources, it is well worth the time to assess the likelihood of an arsenic source being present

within the aquifer matrix. Developing a set of ‘arsenic source risk factors’ according to aquifer depositional history and lithology, may indicate areas at high risk of arsenic contamination prior to fieldwork, which can then be targeted in groundwater surveys. The lessons learned from a small coastal sedimentary aquifer in Australia should be utilised in the assessment and identifi-cation of potential arsenic-rich groundwaters of Latin America.

6 CONCLUSIONS

At Stuarts Point, mobilisation of arsenic occurs when iron oxide surfaces coating sediment grains are dissolved under reducing conditions, releasing their adsorbed arsenic into the groundwater. Increased reducing conditions develop with depth in the aquifer, as do concentra-tions of arsenic. Anoxic conditions in the aquifer develop due to decreased influx of oxidized rainwater, water-sediment interaction and increased flux of organic matter. Progression of the geochemical conditions into a strongly reducing environment promotes precipitation of arsenic into iron sulfides, with sulfate being supplied by seawater intrusion into the aquifer due to ex-cessive exploitation of the groundwater resource. This removes arsenic from the groundwater and incorporates it into a solid phase sink, preventing further migration of dissolved arsenic.

The low risk of anthropogenic arsenic input in combination with the accepted geomorphic model for the Stuarts Point aquifer indicated the aquifer was not expected to contain elevated concentrations of dissolved arsenic in the matrix. Characterisation of the arsenic geochemistry of the Stuarts Point aquifer can be applied as a potential risk assessment tool for fluvial and coastal aquifers of Latin America. A geomorphic approach to aquifer characterisation and iden-tification of potential arsenic sources may indicate Latin American aquifers that could be sus-ceptible to elevated arsenic concentrations. In addition, an assessment of hydrogeochemical conditions to determine the probability of arsenic mobilisation in the aquifer, can aid in the placement of potentially ‘safe’ wells for human consumption of groundwater. Heterogeneity of the aquifer chemical conditions and anthropogenic use of the aquifer should also be considered in the determination of low arsenic groundwater environments.

7 ACKNOWLEDGEMENTS

The authors would like to thank the New South Wales Department of Infrastructure, Planning and Natural Resources for financial assistance with this study. Field and laboratory assistance was provided by Sarah Groves, John Wischusen, Irene Wainwright and Dorothy Yu.

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APPENDIX B

Raw Data

B1 Sediment Chemistry

B2 Groundwater Chemistry

B3 Grain Size Analysis

B4 Sediment Mineral Composition

B5 Iron Oxide Conversions

B6 Sequential Extraction Results

B7 Statistical Results

B8 Geochemical Modelling Results

B9 Selected X-Y plots

(for sediments & groundwater)

Sam

ple

LOI (

%)

SiO

2Ti

O2

Al 2O

3Fe

2O3

MnO

MgO

CaO

Na 2

OK

2OP 2

O5

SO3

Num

ber

Det

ectio

n Li

mit

61.5

27.7

25.6

11.7

13.7

59.7

12.4

164.

29.

712

.620

.4

FAR

M S

OIL

5.05

92.8

10.

056

0.25

0.2

0.01

30.

370.

400.

30.

091

0.08

70.

089

CO

FFE

E R

OC

K1.

5195

.01

0.08

11.

240.

60.

005

0.23

0.03

0.36

0.29

90.

107

0.04

4

ML7

/1.5

0.6

98.6

30.

060.

760.

26nd

0.26

nd0.

480.

170.

040.

04M

L7/4

.50.

7898

.68

0.07

1.44

0.35

0.01

0.2

nd0.

430.

470.

040.

02M

L7/8

3.79

87.4

60.

091.

470.

530.

010.

454.

860.

520.

660.

040.

16M

L7/1

21.

8192

.76

0.08

1.20

0.44

0.01

0.29

2.43

0.45

0.44

0.04

0.09

ML7

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54.

3187

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0.12

1.98

2.02

0.02

0.54

6.02

0.7

0.86

0.05

0.4

ML7

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4.25

83.6

40.

236

3.64

1.1

0.01

50.

465.

450.

650.

715

0.05

20.

582

ML7

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54.

3181

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0.26

3.56

1.17

0.02

0.54

6.02

0.7

0.86

0.05

0.4

ML7

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54.

2370

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0.61

12.2

93.

280.

041.

141.

811.

231.

910.

082.

29M

L7/2

76.

1977

.89

0.24

3.50

1.52

0.02

0.44

5.41

0.58

0.67

0.04

1.93

ML8

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1.96

95.1

0.9

1.35

0.31

0.01

0.19

0.02

0.54

0.31

0.04

0.06

ML8

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0.33

96.6

30.

070.

870.

250.

010.

170.

030.

330.

290.

030.

02M

L8/1

2.5

1.88

90.7

40.

071.

130.

490.

010.

33.

010.

470.

440.

040.

09M

L8/1

52.

1392

.79

0.06

0.84

0.47

0.01

0.29

2.51

0.46

0.27

0.04

0.06

ML8

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2.38

83.7

40.

267.

291.

250.

020.

590.

940.

921.

750.

040.

07M

L8/2

11.

4579

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4610

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1.43

0.02

0.6

0.38

1.1

2.18

0.03

0.48

ML8

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1.18

89.6

60.

162.

220.

830.

010.

312.

300.

720.

650.

040.

75M

L8/3

0.5

5.34

70.0

80.

7217

.40

1.14

0.01

0.56

0.79

0.42

2.56

0.05

0.06

ML9

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391

.49

0.1

1.29

0.32

0.01

0.17

0.01

0.28

0.21

0.03

0.05

ML9

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1.48

97.4

30.

070.

580.

15nd

0.15

0.00

0.27

0.14

0.03

0.02

ML9

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0.53

98.4

30.

060.

730.

170.

010.

150.

000.

310.

230.

030.

02M

L9/5

.60.

3396

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0.07

40.

900.

230.

005

0.19

0.00

0.36

0.31

60.

030.

044

ML9

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61.

4494

.26

0.07

1.22

0.41

0.01

0.21

1.53

0.38

0.41

0.04

0.06

ML9

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61.

6891

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0.15

2.30

0.8

0.01

0.36

1.77

0.52

0.67

0.04

0.27

ML9

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12.

1485

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2.3

7.34

1.26

0.02

0.48

0.13

0.63

1.62

0.03

0.02

ML9

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61.

2788

.52

0.8

4.84

2.18

0.01

0.37

0.21

0.76

1.24

0.08

0.07

ML9

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14.

1870

.05

0.01

13.8

05.

90.

031.

40.

191.

712.

270.

10.

01

ML1

0/2.

80.

696

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0.36

1.27

0.37

0.01

0.16

0.01

0.33

0.32

0.03

0.07

ML1

0/5.

80.

4296

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1.77

1.73

0.44

0.01

0.19

0.03

0.36

0.5

0.03

0.03

Maj

ors

and

Min

ors

as O

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%)

App

endi

x B

1 - S

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hem

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y

Sam

ple

LOI (

%)

SiO

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Al 2O

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MnO

MgO

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Num

ber

Det

ectio

n Li

mit

61.5

27.7

25.6

11.7

13.7

59.7

12.4

164.

29.

712

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.4

Maj

ors

and

Min

ors

as O

xide

s (e

xpre

ssed

as

%)

ML1

0/13

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0381

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0.52

6.18

1.27

0.02

0.54

3.63

1.31

1.5

0.04

0.17

ML1

0/16

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8684

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0.67

7.04

1.2

0.02

0.52

1.11

1.25

1.52

0.04

0.18

ML1

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.81.

488

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0.04

5.47

0.95

0.02

0.49

0.26

1.19

1.26

0.04

0.05

ML1

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3182

.27

0.27

9.13

2.19

0.02

0.57

0.26

0.71

1.17

0.04

0.76

ML1

0/25

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4187

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0.23

5.82

2.52

0.02

0.42

0.30

0.37

0.5

0.04

0.13

ML1

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5578

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188

6.00

9.05

0.01

60.

360.

340.

30.

390.

257

0.01

7

QA

/QC

SA

MPL

ESM

L8/1

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9695

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91.

350.

310.

010.

190.

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540.

310.

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06

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Los

s on

Igni

tion

na =

not

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lyse

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App

endi

x B

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hem

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Sam

ple

SbSn

Cd

Mo

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SrU

Rb

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As

Ga

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eV

Ba

Num

ber

2.9

3.9

3.0

1.8

1.0

1.0

1.0

0.9

2.9

1.0

2.8

2.0

2.5

1.0

1.8

2.0

2.0

2.9

2.9

15.4

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Farm

soi

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31.

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216

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42.

223

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37.

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02.

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05.

38.

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86.

98.

21.

00.

02.

431

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.8

ML7

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10.6

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1.8

45.3

0.0

11.2

2.7

17.7

0.0

4.3

4.3

2.1

4.5

7.5

2.1

6.5

0.3

38.7

18.9

0.0

ML7

/4.5

10.2

ndnd

0.6

2.3

46.7

0.9

19.3

0.6

26.9

0.0

4.6

2.8

2.3

6.0

8.1

3.3

10.9

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35.1

23.4

72.3

ML7

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6nd

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02.

550

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622

6.6

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31.6

0.7

5.2

6.2

2.6

7.3

9.2

5.6

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37.7

24.1

64.1

ML7

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0.7

2.2

47.3

0.6

116.

20.

028

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85.

35.

82.

18.

59.

36.

30.

41.

629

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L7/1

6.5

6.5

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2.4

2.7

75.4

3.1

146.

72.

033

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67.

24.

43.

311

.311

.113

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118

.334

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L7/1

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4.5

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20.

937

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06.

810

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914

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65.

124

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9.8

ML7

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54.

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14.

518

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3.4

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52.

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522

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ML7

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55.

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59.

120

5.9

15.9

124.

55.

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113

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014

.456

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870

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1.0

394.

0M

L7/2

76.

4nd

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54.

611

2.9

5.2

176.

93.

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57.

77.

45.

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06.

923

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2.4

ML8

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11.4

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0.4

1.9

45.7

0.2

15.5

0.0

21.2

0.5

5.0

4.8

2.6

5.9

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2.0

0.0

0.1

33.5

23.2

18.8

ML8

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11.3

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0.9

2.4

47.7

0.7

16.1

0.0

22.6

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4.2

2.5

2.4

5.6

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2.4

14.9

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31.4

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36.9

ML8

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56.

2nd

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811

0.8

1.9

16.5

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3.3

5.7

2.8

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106.

89.

522

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L8/1

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62.

342

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0.6

12.0

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2.9

5.8

2.9

5.9

7.8

2.0

104.

05.

825

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.0M

L8/1

94.

6nd

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1.2

3.6

85.9

10.2

88.5

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60.5

3.6

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5.9

9.1

37.5

13.0

11.1

8.2

25.3

24.7

60.9

359.

7M

L8/2

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3nd

1.4

0.7

5.0

213.

314

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911

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711

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ML8

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2.1

178.

76.

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3.2

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22.9

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4.8

7.7

4.3

15.4

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113.

415

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ML8

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54.

7nd

2.2

0.8

10.5

178.

355

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13.5

22.9

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18.8

29.5

64.6

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82.8

89.6

216.

484

9.2

ML9

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9nd

0.6

0.1

1.3

33.1

3.7

15.0

2.0

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1.4

3.7

6.6

3.2

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2.4

19.4

10.6

26.3

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24.4

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00.

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932

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613

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81.

33.

52.

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66.

75.

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26.

46.

323

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0nd

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32.3

3.2

15.6

1.5

9.8

2.4

2.6

3.4

2.6

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7.4

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34.8

4.6

25.9

12.1

25.6

ML9

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9.9

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2.2

47.1

0.3

16.6

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23.6

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4.4

3.1

2.1

5.9

8.3

3.1

16.4

0.0

31.6

20.1

30.6

ML9

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63.

5nd

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52.

138

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73.

14.

92.

96.

06.

13.

090

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819

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0.9

1.8

90.4

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105.

72.

325

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24.

310

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111

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45.

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7.8

ML9

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13.

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48.

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62.

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33.

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12.

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7

ML1

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84.

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3.1

4.7

2.6

6.6

40.4

0.0

108.

119

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L10/

5.8

6.0

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0.8

2.4

63.7

5.4

21.9

2.1

18.4

2.1

3.7

4.9

3.2

7.7

5.6

0.9

106.

19.

818

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13.3

4.8

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3.4

76.2

8.2

60.6

2.9

41.7

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6.8

24.6

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1.9

111.

114

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9.0

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9nd

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34.

310

9.3

12.9

95.3

1.8

52.5

5.4

7.7

7.8

8.1

32.2

8.4

10.5

26.5

28.2

36.4

56.9

338.

7M

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17.8

3.0

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0.4

3.3

145.

312

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1.2

1.9

52.6

4.9

7.3

6.9

8.1

29.2

8.4

11.1

18.7

28.4

27.0

52.2

312.

2M

L10/

22.3

3.6

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0.7

5.2

155.

915

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859

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710

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53.

382

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130

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08.

34.

57.

723

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7.7

Trac

e El

emen

ts (m

g kg

-1)

App

endi

x B

1 - S

edim

ent C

hem

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y

Sam

ple

SbSn

Cd

Mo

Nb

ZrY

SrU

Rb

ThPb

As

Ga

ZnC

uN

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oC

rC

eV

Ba

Num

ber

2.9

3.9

3.0

1.8

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1.0

0.9

2.9

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2.8

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2.5

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2.9

15.4

2.9

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Trac

e El

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g kg

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ML1

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5nd

3.7

1.7

5.1

47.7

23.1

42.2

2.3

21.9

4.9

25.7

14.0

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56.3

46.2

20.6

15.7

107.

710

8.3

107.

520

8.0

QA

/QC

SA

MPL

ESM

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26.8

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1.9

3.6

47.0

19.1

30.7

4.4

22.5

5.2

24.3

12.9

7.9

59.1

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80.6

91.2

180.

5M

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747

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33.6

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Num

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Sam

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EC

Eh

D

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Tem

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Num

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Vm

g/L

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g/L

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Lm

g/L

mg/

Lm

g/L

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Lm

g/L

mg/

Lm

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mg/

Lm

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L%

ES

TUA

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116

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10.

010.

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500.

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ES

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0.8

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647.

315

62.4

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0.00

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1.30

2.46

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4E

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AR

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31,0

5023

5.96

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264

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432

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3.52

0.63

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3M

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385

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4.40

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377

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381

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3.96

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432

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316

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4320

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30.5

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5M

L1/2

6M

L1/2

7M

L1/2

8M

L1/2

9M

L1/3

0

ML2

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L2/6

ML2

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L2/8

ML2

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0

Ge

Ho

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LuM

nM

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bN

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App

endi

x B

2 - G

roun

dwat

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hem

istr

y

Sam

ple

Num

ber

ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML3

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L3/4

ML3

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L3/6

ML3

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L3/8

ML3

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L3/1

0M

L3/1

1M

L3/1

2M

L3/1

3M

L3/1

4M

L3/1

5M

L3/1

6M

L3/1

7M

L3/1

8M

L3/1

9

Ge

Ho

LaLi

LuM

nM

oN

bN

dN

iP

PbPr

PtR

bSb

Scug

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/Lug

/Lug

/Lug

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/Lug

/Lug

/Lug

/Lug

/Lug

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App

endi

x B

2 - G

roun

dwat

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hem

istr

y

Sam

ple

Num

ber

ML3

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ML3

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ML3

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ML3

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ML3

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ML3

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ML3

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ML3

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ML4

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L4/4

ML4

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L4/6

ML4

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L4/8

ML4

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L4/1

0M

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L4/1

2M

L4/1

3M

L4/1

4M

L4/1

5M

L4/1

6M

L4/1

7M

L4/1

8M

L4/1

9M

L4/2

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L4/2

2M

L4/2

3M

L4/2

4M

L4/2

5M

L4/2

6M

L4/2

7M

L4/2

8

Ge

Ho

LaLi

LuM

nM

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PbPr

PtR

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Scug

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App

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0.07

0.00

0.08

1.39

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030.

650.

13<0

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0.08

1.12

80.5

23.

580.

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1.68

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511.

860.

100.

010.

160.

97<0

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3.98

0.15

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900.

180.

9912

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0.04

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2.21

0.10

0.02

0.46

2.34

0.00

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60.

20<0

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0.50

0.68

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160.

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320.

120.

020.

603.

050.

0016

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0.19

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900.

680.

6621

1.05

1.04

0.14

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062.

82<0

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2.43

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0.01

0.13

3.41

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0.08

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900.

130.

5920

8.86

0.87

0.03

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2.31

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120.

010.

2715

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0.00

36.5

20.

26<0

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0.32

0.77

476.

934.

080.

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131.

4578

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4.48

0.03

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50.

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1.01

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0.04

0.04

0.25

0.62

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30.

20<0

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0.60

0.99

194.

369.

970.

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890.

020.

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090.

530.

0124

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20.

191.

1614

2.57

4.89

0.03

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061.

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3.15

0.04

0.02

0.14

0.67

0.01

23.0

00.

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0.30

0.98

149.

255.

430.

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2.38

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200.

030.

020.

100.

850.

0120

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0.22

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20.

210.

7212

8.30

0.81

0.04

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061.

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2.82

0.02

0.02

0.16

0.83

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20.

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0.30

1.08

110.

8111

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2.87

0.02

0.01

0.06

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0.21

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20.

111.

0913

6.20

6.72

0.02

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061.

21<0

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2.79

0.06

0.01

0.05

0.94

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0.30

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20.

081.

2016

4.08

2.53

0.02

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061.

07<0

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2.80

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60.

40<0

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0.06

1.16

211.

375.

390.

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010.

050.

010.

061.

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51.6

30.

96<0

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1.35

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153.

650.

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760.

040.

000.

041.

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10.

85<0

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0.07

1.49

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925.

000.

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1.70

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010.

02<0

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0.03

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0.84

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061.

3716

9.85

7.28

0.01

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061.

78<0

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80.

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1.19

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263.

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22.

940.

020.

000.

031.

87<0

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55.9

41.

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152.

824.

060.

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23.

11

App

endi

x B

2 - G

roun

dwat

er C

hem

istr

y

Sam

ple

Num

ber

ML7

/22

ML7

/23

ML7

/24

ML7

/25

ML7

/26

ML7

/27

ML7

/28

ML8

/6M

L8/7

ML8

/8M

L8/9

ML8

/10

ML8

/11

ML8

/12

ML8

/13

ML8

/14

ML8

/15

ML8

/16

ML8

/17

ML8

/18

ML8

/19

ML8

/20

ML8

/21

ML8

/22

ML8

/23

ML8

/24

ML8

/25

ML8

/26

ML8

/27

ML8

/28

ML8

/29

ML9

/5.5

ML9

/6.5

Ge

Ho

LaLi

LuM

nM

oN

bN

dN

iP

PbPr

PtR

bSb

Scug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

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020.

000.

042.

01<0

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60.1

11.

03<0

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0.07

1.47

179.

019.

310.

01<0

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2.67

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23.

340.

010.

010.

043.

040.

0085

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0.96

<0.0

20.

091.

7724

9.66

7.35

0.02

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065.

86<0

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5.35

<0.0

060.

000.

032.

96<0

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52.2

50.

75<0

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0.05

1.75

409.

751.

950.

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6.02

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26.

260.

01<0

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0.02

5.73

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0319

6.65

0.66

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20.

031.

7744

7.85

0.37

0.00

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064.

96<0

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6.34

0.02

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030.

015.

90<0

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156.

460.

74<0

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0.02

1.58

487.

376.

80<0

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064.

54<0

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6.56

0.02

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030.

005.

71<0

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140.

840.

70<0

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0.00

1.96

389.

373.

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064.

48<0

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6.80

0.05

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030.

005.

61<0

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132.

260.

61<0

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0.01

1.66

290.

671.

67<0

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064.

60<0

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6.68

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060.

020.

170.

200.

0040

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0.12

0.03

0.33

0.87

39.4

59.

220.

06<0

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0.82

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20.

290.

010.

010.

050.

26<0

.003

78.8

10.

09<0

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0.10

0.95

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81.

540.

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20.

330.

050.

010.

070.

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96.0

60.

43<0

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0.14

1.29

7.24

4.50

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061.

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0.52

0.04

0.01

0.09

0.27

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0382

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0.42

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20.

181.

46<6

.00

4.41

0.03

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061.

39<0

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0.27

0.06

0.01

0.07

0.26

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0365

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0.79

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20.

151.

1435

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2.86

0.03

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061.

780.

160.

570.

030.

010.

120.

13<0

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42.1

50.

31<0

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0.20

0.92

14.5

12.

600.

04<0

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1.07

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2<0

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0.03

0.01

0.10

0.14

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0324

.81

0.21

<0.0

20.

180.

6148

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5.87

0.04

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060.

78<0

.02

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20.

040.

010.

120.

13<0

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14.9

80.

19<0

.02

0.21

0.38

93.2

92.

450.

04<0

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0.67

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0.03

0.01

0.14

0.13

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038.

260.

22<0

.02

0.21

0.35

92.9

43.

100.

05<0

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0.87

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20.

120.

020.

000.

050.

11<0

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10.4

80.

31<0

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0.09

0.48

87.0

65.

810.

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0.78

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2<0

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0.01

0.01

0.08

0.15

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0319

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0.20

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20.

130.

6884

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8.68

0.03

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061.

28<0

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0.29

0.02

0.01

0.06

0.20

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0.20

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20.

110.

8350

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15.2

50.

02<0

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1.62

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20.

240.

010.

000.

040.

23<0

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42.5

20.

24<0

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0.07

0.72

55.1

415

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0.01

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061.

44<0

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0.43

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06<0

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0.03

0.58

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0370

.10

0.13

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20.

050.

9248

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1.03

0.01

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061.

80<0

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0.60

0.01

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030.

020.

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80.3

10.

27<0

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0.03

0.94

23.8

03.

550.

01<0

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2.02

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20.

690.

01<0

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0.01

0.33

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0384

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0.14

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20.

020.

909.

789.

770.

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1.16

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20.

500.

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0.01

0.28

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0386

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0.13

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20.

011.

079.

512.

590.

00<0

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1.23

<0.0

20.

670.

02<0

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0.01

0.18

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0385

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0.23

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20.

011.

15<6

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11.1

10.

00<0

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0.95

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20.

660.

02<0

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0.01

0.25

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0385

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0.18

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20.

011.

58<6

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4.12

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1.20

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20.

670.

02<0

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0.01

0.20

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0381

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0.14

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20.

010.

91<6

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2.61

0.00

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061.

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0.59

0.02

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010.

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83.8

00.

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0.02

1.10

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05.

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00<0

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1.05

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20.

560.

02<0

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0.01

1.22

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0.17

0.03

0.02

1.43

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08.

260.

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1.86

0.04

1.46

0.05

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030.

011.

34<0

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21.8

80.

290.

020.

021.

3923

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7.46

0.00

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061.

93<0

.02

2.49

0.03

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030.

012.

82<0

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35.6

71.

16<0

.02

0.01

1.94

66.7

713

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0.00

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062.

45<0

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2.73

0.02

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030.

041.

30<0

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40.6

00.

11<0

.02

0.06

1.64

172.

022.

360.

01<0

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0.61

<0.0

21.

200.

010.

010.

051.

16<0

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76.7

80.

10<0

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0.08

0.88

118.

481.

840.

02<0

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0.78

<0.0

20.

83

App

endi

x B

2 - G

roun

dwat

er C

hem

istr

y

Sam

ple

Num

ber

ML9

/7.5

ML9

/8.5

ML9

/9.5

ML9

/10.

5M

L9/1

2.5

ML9

/13.

5M

L9/1

4.5

ML9

/15.

5M

L9/1

6.5

ML9

/17.

5M

L9/1

8.5

ML9

/19.

5M

L9/2

0.5

ML9

/21.

5

ML1

0/5

ML1

0/6

ML1

0/7

ML1

0/9

ML1

0/10

ML1

0/11

ML1

0/12

ML1

0/13

ML1

0/16

ML1

0/21

ML1

0/22

ML1

0/23

Ge

Ho

LaLi

LuM

nM

oN

bN

dN

iP

PbPr

PtR

bSb

Scug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/Lug

/L0.

010.

010.

101.

83<0

.003

82.9

30.

160.

020.

171.

3888

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1.64

0.03

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061.

61<0

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1.36

0.01

0.01

0.12

0.73

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0392

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0.07

0.02

0.20

0.91

69.0

50.

890.

04<0

.006

1.93

<0.0

21.

220.

010.

020.

120.

980.

0190

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0.07

<0.0

20.

220.

7711

9.59

3.41

0.04

<0.0

062.

58<0

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1.37

0.02

0.01

0.07

0.83

0.00

51.5

60.

11<0

.02

0.12

0.70

128.

214.

970.

02<0

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2.49

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21.

080.

010.

010.

070.

54<0

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65.5

50.

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0.11

1.09

74.5

23.

050.

02<0

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2.81

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21.

130.

020.

010.

070.

64<0

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66.5

00.

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0.11

0.96

87.1

61.

800.

02<0

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3.32

<0.0

21.

270.

020.

010.

060.

66<0

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85.2

30.

06<0

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0.10

1.28

70.7

42.

850.

02<0

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3.43

<0.0

21.

190.

020.

010.

051.

15<0

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78.7

40.

06<0

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0.09

1.05

111.

301.

370.

02<0

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3.47

<0.0

21.

290.

020.

000.

062.

00<0

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68.6

60.

22<0

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0.09

1.01

125.

571.

740.

02<0

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3.64

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21.

450.

020.

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062.

82<0

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64.4

70.

07<0

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0.08

1.12

138.

825.

030.

02<0

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3.08

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21.

470.

02<0

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0.03

1.39

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0383

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0.10

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20.

052.

9514

5.83

3.30

0.01

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062.

37<0

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1.42

0.04

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030.

043.

61<0

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93.9

80.

13<0

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0.04

1.27

176.

056.

980.

01<0

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2.48

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21.

940.

06<0

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0.01

2.76

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0312

0.24

0.31

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20.

011.

4620

7.90

2.97

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3.97

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22.

270.

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033.

35<0

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166.

450.

640.

040.

011.

1632

9.00

2.08

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03<0

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4.35

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22.

30

0.13

0.01

0.05

0.39

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2.29

4.55

0.05

0.08

7.14

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01.

530.

02<0

.006

7.07

1.28

1.42

0.13

0.01

0.22

0.37

0.00

599.

452.

890.

020.

322.

83<6

.00

1.85

0.06

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065.

250.

031.

250.

040.

010.

120.

450.

0117

3.98

2.63

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20.

152.

2011

.45

4.66

0.03

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064.

17<0

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1.15

0.04

0.01

0.10

0.37

0.01

184.

372.

56<0

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0.15

1.90

15.0

81.

470.

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3.30

<0.0

21.

080.

040.

010.

070.

360.

0114

6.69

9.04

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20.

112.

247.

564.

700.

02<0

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3.37

<0.0

20.

970.

060.

010.

080.

470.

0018

4.15

7.01

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20.

121.

99<6

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2.89

0.03

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063.

43<0

.02

1.06

0.09

0.01

0.08

0.47

0.00

208.

793.

41<0

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0.13

1.90

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021

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0.03

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063.

17<0

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0.95

0.14

0.01

0.07

0.58

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0326

9.74

2.37

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20.

121.

81<6

.00

6.44

0.02

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063.

290.

031.

230.

170.

010.

030.

68<0

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223.

003.

30<0

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0.05

1.46

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77

App

endi

x B

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hem

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Sam

ple

Num

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1E

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AR

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3

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L1/6

ML1

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L1/8

ML1

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L1/1

0M

L1/1

1M

L1/1

2M

L1/1

3M

L1/1

4M

L1/1

5M

L1/1

6M

L1/1

7M

L1/1

8M

L1/1

9M

L1/2

0M

L1/2

1M

L1/2

4M

L1/2

5M

L1/2

6M

L1/2

7M

L1/2

8M

L1/2

9M

L1/3

0

ML2

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L2/6

ML2

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L2/8

ML2

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0

SeSm

SnSr

TaTb

ThU

VW

YYb

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ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

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75.2

20.

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2873

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0.40

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0.28

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1939

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220.

0113

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090.

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6105

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0.12

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07

0.33

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3930

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0.20

0.00

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0.88

0.34

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080.

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80.

160.

010.

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550.

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330.

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270.

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120.

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2.15

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0.51

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980.

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0.07

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3937

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0.04

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980.

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3933

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0.30

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980.

22<0

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3926

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980.

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0.10

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3928

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0.46

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980.

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4.30

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0.36

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1553

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0.02

0.01

0.99

0.16

0.64

0.05

1.12

0.07

0.33

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210.

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0.06

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2.52

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25

App

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Sam

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Num

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

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ML2

/19

ML2

/20

ML2

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ML2

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ML2

/23

ML2

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ML2

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ML2

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ML2

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ML3

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L3/4

ML3

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L3/6

ML3

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L3/8

ML3

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L3/1

0M

L3/1

1M

L3/1

2M

L3/1

3M

L3/1

4M

L3/1

5M

L3/1

6M

L3/1

7M

L3/1

8M

L3/1

9

SeSm

SnSr

TaTb

ThU

VW

YYb

ZnZr

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

0.36

0.04

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3944

9.47

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0.02

2.43

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510.

240.

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0.34

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445.

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040.

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56<0

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3938

6.85

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0.02

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2.42

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510.

190.

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0.42

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090.

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577.

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6.59

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0.38

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App

endi

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dwat

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hem

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Sam

ple

Num

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ML3

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ML4

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L4/6

ML4

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L4/8

ML4

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L4/1

0M

L4/1

1M

L4/1

2M

L4/1

3M

L4/1

4M

L4/1

5M

L4/1

6M

L4/1

7M

L4/1

8M

L4/1

9M

L4/2

0M

L4/2

1M

L4/2

2M

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3M

L4/2

4M

L4/2

5M

L4/2

6M

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7M

L4/2

8

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SnSr

TaTb

ThU

VW

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ug/L

ug/L

ug/L

ug/L

ug/L

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ug/L

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Sam

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ML4

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L8/9

ML8

/10

ML8

/11

ML8

/12

ML8

/13

ML8

/14

ML8

/15

ML8

/16

ML8

/17

ML8

/18

ML8

/19

ML8

/20

ML8

/21

ML8

/22

ML8

/23

ML8

/24

ML8

/25

ML8

/26

ML8

/27

ML8

/28

ML8

/29

ML9

/5.5

ML9

/6.5

SeSm

SnSr

TaTb

ThU

VW

YYb

ZnZr

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

0.94

0.02

<0.1

837

9.78

<0.0

15<0

.003

<0.0

060.

251.

250.

100.

110.

01<0

.42

0.18

2.79

0.02

<0.1

853

4.23

<0.0

150.

00<0

.006

0.90

3.29

0.12

0.15

0.02

<0.4

20.

163.

030.

01<0

.18

500.

92<0

.015

<0.0

03<0

.006

0.78

3.36

0.09

0.09

0.01

<0.4

20.

263.

070.

01<0

.18

484.

09<0

.015

<0.0

03<0

.006

0.50

3.46

0.08

0.05

0.01

<0.4

20.

303.

120.

01<0

.18

473.

53<0

.015

<0.0

03<0

.006

0.39

3.29

0.07

0.04

0.00

<0.4

20.

293.

14<0

.003

<0.1

846

9.28

<0.0

15<0

.003

<0.0

060.

223.

280.

070.

02<0

.003

4.82

0.26

2.95

0.00

<0.1

846

9.06

<0.0

15<0

.003

<0.0

060.

163.

090.

050.

020.

0013

.92

0.22

<0.2

40.

08<0

.18

341.

41<0

.015

0.01

0.04

0.09

1.62

0.05

0.50

0.03

<0.4

20.

97<0

.24

0.03

<0.1

844

5.04

<0.0

150.

01<0

.006

0.10

<0.1

0<0

.018

0.17

0.01

<0.4

20.

420.

560.

03<0

.18

623.

84<0

.015

0.01

0.01

<0.0

030.

19<0

.018

0.23

0.01

<0.4

20.

530.

580.

04<0

.18

508.

55<0

.015

0.01

0.02

<0.0

030.

43<0

.018

0.31

0.02

<0.4

20.

540.

450.

03<0

.18

532.

37<0

.015

0.01

<0.0

06<0

.003

0.24

0.02

0.25

0.01

<0.4

20.

550.

320.

04<0

.18

368.

20<0

.015

0.01

0.01

<0.0

03<0

.10

0.05

0.22

0.01

<0.4

20.

50<0

.24

0.04

<0.1

826

7.01

<0.0

150.

01<0

.006

<0.0

030.

130.

050.

250.

01<0

.42

0.57

<0.2

40.

04<0

.18

211.

42<0

.015

0.01

<0.0

060.

06<0

.10

0.06

0.23

0.01

<0.4

20.

42<0

.24

0.04

<0.1

823

9.65

<0.0

150.

01<0

.006

<0.0

030.

160.

060.

200.

01<0

.42

0.68

<0.2

40.

02<0

.18

233.

49<0

.015

0.00

<0.0

06<0

.003

<0.1

00.

050.

120.

015.

200.

47<0

.24

0.03

<0.1

833

3.13

<0.0

150.

00<0

.006

<0.0

03<0

.10

0.04

0.15

0.00

<0.4

20.

36<0

.24

0.03

<0.1

832

8.89

<0.0

150.

01<0

.006

<0.0

03<0

.10

0.04

0.17

0.01

<0.4

20.

43<0

.24

0.02

<0.1

835

4.66

<0.0

150.

00<0

.006

<0.0

03<0

.10

<0.0

180.

120.

00<0

.42

0.33

<0.2

40.

01<0

.18

441.

90<0

.015

<0.0

03<0

.006

0.01

<0.1

0<0

.018

0.06

<0.0

03<0

.42

0.28

<0.2

40.

01<0

.18

479.

86<0

.015

<0.0

03<0

.006

<0.0

03<0

.10

<0.0

180.

04<0

.003

<0.4

20.

30<0

.24

0.00

<0.1

848

9.34

<0.0

15<0

.003

<0.0

06<0

.003

<0.1

0<0

.018

0.05

<0.0

03<0

.42

0.19

<0.2

40.

00<0

.18

513.

64<0

.015

<0.0

03<0

.006

<0.0

03<0

.10

<0.0

180.

04<0

.003

<0.4

20.

18<0

.24

0.00

<0.1

849

5.22

<0.0

15<0

.003

<0.0

06<0

.003

<0.1

0<0

.018

0.04

<0.0

035.

900.

22<0

.24

<0.0

03<0

.18

502.

55<0

.015

<0.0

03<0

.006

<0.0

03<0

.10

<0.0

180.

03<0

.003

<0.4

20.

21<0

.24

<0.0

03<0

.18

478.

05<0

.015

<0.0

03<0

.006

<0.0

03<0

.10

<0.0

180.

04<0

.003

<0.4

20.

21<0

.24

0.00

<0.1

848

6.77

<0.0

15<0

.003

<0.0

06<0

.003

<0.1

0<0

.018

0.08

<0.0

03<0

.42

0.26

<0.2

40.

00<0

.18

509.

030.

03<0

.003

<0.0

060.

40<0

.10

0.05

0.09

<0.0

03<0

.42

0.29

<0.2

40.

00<0

.18

393.

200.

03<0

.003

0.02

0.12

<0.1

00.

060.

10<0

.003

<0.4

20.

550.

40<0

.003

<0.1

825

6.53

0.02

<0.0

03<0

.006

0.19

<0.1

00.

080.

06<0

.003

1.23

0.08

<0.2

40.

01<0

.18

515.

900.

02<0

.003

0.02

<0.0

03<0

.10

0.08

0.08

0.01

<0.4

20.

25<0

.24

0.02

<0.1

846

6.43

0.02

0.00

0.04

<0.0

030.

240.

050.

150.

01<0

.42

0.36

App

endi

x B

2 - G

roun

dwat

er C

hem

istr

y

Sam

ple

Num

ber

ML9

/7.5

ML9

/8.5

ML9

/9.5

ML9

/10.

5M

L9/1

2.5

ML9

/13.

5M

L9/1

4.5

ML9

/15.

5M

L9/1

6.5

ML9

/17.

5M

L9/1

8.5

ML9

/19.

5M

L9/2

0.5

ML9

/21.

5

ML1

0/5

ML1

0/6

ML1

0/7

ML1

0/9

ML1

0/10

ML1

0/11

ML1

0/12

ML1

0/13

ML1

0/16

ML1

0/21

ML1

0/22

ML1

0/23

SeSm

SnSr

TaTb

ThU

VW

YYb

ZnZr

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

ug/L

<0.2

40.

04<0

.18

615.

55<0

.015

0.01

0.08

<0.0

031.

330.

020.

280.

021.

790.

40<0

.24

0.05

<0.1

848

7.26

<0.0

150.

010.

06<0

.003

0.91

0.02

0.28

0.02

<0.4

20.

39<0

.24

0.06

<0.1

839

7.48

<0.0

150.

010.

06<0

.003

1.81

0.03

0.49

0.05

<0.4

20.

60<0

.24

0.03

<0.1

841

3.70

<0.0

150.

010.

03<0

.003

1.80

0.02

0.26

0.03

<0.4

20.

45<0

.24

0.03

<0.1

850

4.82

<0.0

150.

010.

03<0

.003

0.81

0.03

0.19

0.02

<0.4

20.

78<0

.24

0.02

<0.1

855

4.83

<0.0

150.

000.

02<0

.003

0.74

0.02

0.18

0.01

<0.4

20.

70<0

.24

0.02

<0.1

857

8.08

<0.0

150.

000.

02<0

.003

0.43

0.02

0.16

0.01

<0.4

20.

360.

250.

02<0

.18

608.

47<0

.015

0.00

0.01

<0.0

030.

460.

020.

150.

01<0

.42

0.53

<0.2

40.

02<0

.18

595.

10<0

.015

0.00

0.02

<0.0

030.

610.

020.

150.

01<0

.42

1.34

<0.2

40.

02<0

.18

658.

87<0

.015

<0.0

03<0

.006

<0.0

030.

25<0

.018

0.12

0.01

<0.4

20.

38<0

.24

0.02

<0.1

862

3.24

<0.0

15<0

.003

<0.0

06<0

.003

<0.1

00.

040.

110.

01<0

.42

0.24

<0.2

40.

01<0

.18

662.

13<0

.015

<0.0

03<0

.006

<0.0

03<0

.10

0.03

0.10

0.00

<0.4

20.

57<0

.24

0.01

<0.1

869

1.22

<0.0

15<0

.003

<0.0

060.

09<0

.10

0.04

0.05

<0.0

03<0

.42

0.23

<0.2

4<0

.003

<0.1

862

6.04

0.02

<0.0

030.

030.

20<0

.10

0.07

0.04

<0.0

03<0

.42

0.64

1.06

0.02

<0.1

825

90.4

00.

020.

000.

054.

663.

810.

050.

260.

02<0

.42

1.28

0.64

0.06

<0.1

810

52.1

8<0

.015

0.01

0.06

<0.0

030.

390.

030.

350.

03<0

.42

0.56

0.43

0.03

<0.1

876

5.94

<0.0

150.

010.

01<0

.003

<0.1

00.

020.

190.

04<0

.42

0.36

0.28

0.03

<0.1

874

4.86

<0.0

150.

010.

01<0

.003

<0.1

00.

030.

220.

03<0

.42

0.62

0.25

0.03

<0.1

874

6.73

<0.0

150.

000.

01<0

.003

<0.1

00.

030.

240.

04<0

.42

0.12

0.35

0.03

0.39

763.

27<0

.015

0.00

0.01

<0.0

03<0

.10

0.02

0.19

0.02

<0.4

20.

250.

360.

03<0

.18

736.

27<0

.015

0.00

0.01

<0.0

03<0

.10

0.02

0.19

0.02

<0.4

20.

340.

320.

030.

4071

8.60

<0.0

150.

000.

02<0

.003

<0.1

00.

020.

210.

02<0

.42

0.41

0.44

0.02

0.67

592.

39<0

.015

<0.0

03<0

.006

<0.0

03<0

.10

0.02

0.14

0.01

<0.4

20.

390.

800.

034.

3645

8.59

<0.0

150.

01<0

.006

<0.0

030.

230.

050.

270.

02<0

.42

0.27

1.31

0.03

7.80

458.

49<0

.015

0.01

<0.0

06<0

.003

0.41

0.07

0.28

0.02

<0.4

20.

321.

650.

027.

4742

1.72

<0.0

150.

01<0

.006

<0.0

030.

570.

070.

220.

01<0

.42

0.21

App

endi

x B

2 - G

roun

dwat

er C

hem

istr

y

Com

paris

on o

f Gra

phic

al v

ersu

s M

omen

t Mea

sure

s M

etho

d in

Cal

cula

tion

of G

rain

Siz

e Pa

ram

eter

s fo

r San

d Pr

oven

ance

Det

erm

inat

ion

Sam

ple

IDR

elat

ive

Perc

ent D

iffer

ence

(RPD

s)A

llow

able

err

or u

p to

5%

RPD

ML7

/1.5

ML7

/4.5

ML7

/8M

L7/1

2M

L7/1

8m

ean

mea

n G

1.97

1.82

2.08

2.24

1.88

3.88

3.31

4.33

5.02

3.53

4.01

mea

n M

2.16

2.04

2.35

2.50

2.53

std

dev

G0.

380.

500.

630.

680.

530.

140.

250.

400.

460.

280.

31st

d de

v M

0.36

0.48

0.65

0.69

0.87

skew

ness

G-0

.02

-0.1

1-0

.11

0.12

-0.3

50.

000.

010.

010.

010.

120.

03sk

ewne

ss M

0.00

-0.0

10.

070.

09-0

.22

ML8

/1.5

ML8

/4.5

ML8

/12.

5M

L8/1

5M

L8/2

4m

ean

G2.

291.

792.

171.

461.

825.

243.

204.

712.

133.

313.

72m

ean

M2.

512.

022.

401.

742.

10st

d de

v G

0.27

0.47

0.48

0.83

0.77

0.07

0.22

0.23

0.69

0.59

0.36

std

dev

M0.

290.

500.

560.

840.

84sk

ewne

ss G

0.23

-0.0

1-0

.18

0.40

0.15

0.05

0.00

0.03

0.16

0.02

0.05

skew

ness

M0.

000.

01-0

.03

0.46

0.26

ML9

/1M

L9/2

.6M

L9/4

.1M

L9/5

.6M

L9/1

1.6

mea

n G

2.05

1.79

2.25

2.11

2.39

4.20

3.20

5.06

4.45

5.71

4.53

mea

n M

2.28

2.02

2.43

2.34

2.55

std

dev

G0.

330.

460.

310.

400.

470.

110.

210.

100.

160.

220.

16st

d de

v M

0.37

0.46

0.29

0.37

0.51

skew

ness

G-0

.05

-0.0

1-0

.10

-0.1

40.

190.

000.

000.

010.

020.

040.

01sk

ewne

ss M

0.00

0.00

0.00

0.00

0.01

ML1

0/2.

8M

L10/

5.8

coffe

e ro

ckfa

rm s

oil

mea

n G

2.33

2.33

2.21

2.14

5.43

5.43

4.88

4.58

5.08

mea

n M

2.26

2.11

2.09

1.97

std

dev

G0.

360.

500.

440.

470.

130.

250.

190.

220.

20st

d de

v M

0.36

0.50

0.40

0.54

skew

ness

G-0

.28

-0.0

60.

09-0

.02

0.08

0.00

0.01

0.00

0.02

skew

ness

M0.

000.

000.

00-0

.04

G =

gra

phic

al

M =

mom

ent m

easu

res

met

hod

st

d de

v =

stan

dard

dev

iatio

n

App

endi

x B

3 - G

rain

Siz

e A

naly

sis

Estimate of mineral composition for ternary diagram

Sample IDFeldspar

%Quartz

% Lithics

%ML7/1.5 4 96 0ML7/12 4 82 6ML7/18 5 85 5ML7/4.5 8 89 3ML7/8 2 90 5ML8/1.5 3 97 0ML8/12.6 7 83 4ML8/15 10 75 10ML8/24 6 77 9ML8/4.5 6 89 5ML9/1 5 94 1ML9/11.6 5 90 4ML9/2.6 5 93 2ML9/4.1 3 93 4ML9/5.6 3 92 5ML10/2.8 4 93 3ML10/5.8 10 92 8

Appendix B4 - Ternary Diagram

Estimate of total iron attributable to iron oxide - ML7/18

AW %Fe 55.847 111.694 69.94 Fe in Fe2O3O 16.0 48 30.06 O in Fe2O3

159.694 100.00

XRF analysis Fe2O3 100 g of sample contains: 1.10 g Fe2O30.77 g Fe/100g

0.0076937 g Fe/1g7.6936767 mg/g

DCB Extractant 4.6 mg/l0.115 mg/g

Total % of iron attributable to iron oxides = 1.49 %

Estimate of total aluminium attributable to aluminium oxide - ML7/18

AW %Al 26.98 53.96 52.92 Al in Al2O3O 16 48 47.08 O in Al2O3

101.96 100

XRF analysis Al2O3 100 g of sample contains: 3.64 g Al2O31.93 g Al/100g

0.0192639 g Al/1g19.263868 mg/g


Total % of aluminium attributable to Al oxides = 0.07 %

Appendix B5 - Oxide Conversions

Estimate of total iron attributable to iron oxide - ML10/26.8


159.694 100.00


0.063298 g Fe/1g63.297976 mg/g



Estimate of total aluminium attributable to aluminium oxide - ML10/26.8


101.96 100


0.0317536 g Al/1g31.753629 mg/g




Estimate of total iron attributable to iron oxide - ML9/5.6


159.694 100.00


0.0016087 g Fe/1g1.6086778 mg/g

DCB Extractant 0 mg/l0.000 mg/g


Estimate of total aluminium attributable to aluminium oxide - ML9/5.6


101.96 100


0.004763 g Al/1g4.7630443 mg/g




Estimate of total iron attributable to iron oxide - coffee rock


159.694 100.00


0.0041966 g Fe/1g4.1965509 mg/g



Estimate of total aluminium attributable to aluminium oxide - coffee rock


101.96 100


0.0065624 g Al/1g6.5624166 mg/g




Estimate of total iron attributable to iron oxide - farm topsoil


159.694 100.00


0.0013989 g Fe/1g1.3988503 mg/g



Estimate of total aluminium attributable to aluminium oxide - farm topsoil


101.96 100


0.0013231 g Al/1g1.3230679 mg/g




Uni

vers

ity o

f N.S

.W.

Dat

e A

naly

sed

21st

May

200

3S

choo

l of B

iolo

gica

l, E

arth

and

Env

ironm

enta

l Sci

ence

sFi

le A

ddre

ssc:

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lysi

s\re

port\

2003

\Bet

h O

'She

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Dendrogram distinguishing groundwater groups at Stuarts Point

C A S E 0 5 10 15 20 25 Sample # +---------+---------+---------+---------+---------+ ML8/22 40 òø ML8/26 44 òú ML8/23 41 òú ML1/17 88 òú ML3/9 129 òú ML3/10 130 òú ML8/9 27 òú ML1/18 89 òú ML8/24 42 òú ML9/8.5 51 òú ML3/4 124 òú ML8/27 45 òú ML9/5.5 48 òú ML9/12.5 54 òú ML7/17 12 òú ML7/19 14 òú ML7/20 15 òú ML7/18 13 òú ML7/21 16 òú ML5/15 187 òú ML5/18 190 òú ML5/19 191 òú ML5/16 188 òú ML5/20 192 òú ML7/22 17 òú ML5/17 189 òú ML10/21 73 òú ML10/10 68 òú ML10/11 69 òú ML10/9 67 òú ML10/12 70 òú ML2/17 112 òôòòòòòø ML2/18 113 òú ó ML2/19 114 òú ó ML2/14 109 òú ó ML2/15 110 òú ó ML2/10 105 òú ó ML9/6.5 49 òú ó ML3/11 131 òú ó ML1/20 91 òú ó ML4/11 156 òú ó ML8/10 28 òú ó ML1/16 87 òú ó ML2/16 111 òú ó ML2/20 115 òú ó ML1/19 90 òú ó ML2/22 117 òú ó ML2/21 116 òú ó ML8/21 39 òú ó ML8/25 43 òú ó ML8/7 25 òú ó ML8/20 38 òú ó ML8/28 46 òú ó ML1/21 92 òú ó ML2/23 118 òú ó ML1/15 86 òú ó ML4/8 153 òú ó ML4/9 154 òú ùòòòòòòòòòòòø ML4/10 155 òú ó ó ML2/24 119 òú ó ó ML3/5 125 òú ó ó ML2/25 120 òú ó ó ML2/26 121 òú ó ó ML2/27 122 òú ó ó ML2/11 106 ò÷ ó ó ML5/27 197 òø ó ó ML6/29 223 òú ó ó ML7/28 23 òú ó ó ML1/24 93 òú ó ó ML10/22 74 òú ó ó ML10/23 75 òú ó ó ML7/26 21 òú ó ó ML7/27 22 òú ó ó ML7/25 20 òú ó ó ML7/23 18 òôòø ó ó ML7/24 19 òú ó ó ó ML9/13.5 55 òú ó ó ó ML9/17.5 59 òú ó ó ó ML9/16.5 58 òú ó ó ó ML9/18.5 60 òú ó ó ó ML8/8 26 òú ó ó ó ML10/7 66 òú ó ó ó ML10/13 71 òú ó ó ó ML9/21.5 63 òú ó ó ó ML10/16 72 òú ùòòò÷ ó ML9/7.5 50 òú ó ó ML9/14.5 56 òú ó ó ML9/15.5 57 òú ó ó ML9/19.5 61 òú ó ó ML9/20.5 62 ò÷ ó ó ML5/30 200 òø ó ó ML6/30 224 òú ó ùòòòòòòòòòòòòòòòòòòòòòòòòòòòòòø ML5/28 198 òú ó ó ó ML5/29 199 òú ó ó ó ML10/6 65 òôò÷ ó ó ML1/25 94 òú ó ó ML10/5 64 ò÷ ó ó ML1/7 78 òø ó ó ML4/5 150 òú ó ó ML5/6 178 òú ó ó ML4/3 148 òú ó ó ML4/6 151 òú ó ó ML6/7 201 òú ó ó ML5/4 176 òú ó ó ML1/5 76 òú ó ó ML3/3 123 òôòòòòòòòòòø ó ó ML4/4 149 òú ó ó ó ML1/6 77 òú ó ó ó ML5/5 177 òú ó ó ó ML6/9 203 òú ó ó ó ML6/10 204 òú ó ó ó ML2/5 100 òú ó ó ó ML2/6 101 òú ó ó ó ML2/8 103 òú ó ó ó ML6/8 202 òú ó ó ó ML2/7 102 ò÷ ó ó ó ML5/22 194 òø ó ó ó ML5/26 196 òú ó ó ó ML5/23 195 òú ó ó ó ML8/29 47 òú ó ó ó ML6/27 221 òú ó ó ó ML6/28 222 òú ó ó ó ML6/26 220 òú ó ó ó ML5/21 193 òú ó ó ó ML6/25 219 òú ó ó ó ML3/13 133 òú ó ó ó ML6/20 214 òú ùòòòòòòò÷ ó ML7/9 4 òú ó ó ML3/7 127 òôòø ó ó ML3/12 132 òú ó ó ó ML1/8 79 òú ó ó ó ML5/14 186 òú ó ó ó ML6/23 217 òú ó ó ó ML6/24 218 òú ó ó ó ML8/17 35 òú ó ó ó ML6/21 215 òú ó ó ó ML7/7 2 òú ó ó ó ML7/15 10 òú ó ó ó ML7/16 11 òú ó ó ó ML8/18 36 òú ó ó ó ML7/10 5 òú ó ó ó ML8/6 24 òú ó ó ó ML8/11 29 òú ó ó ó ML3/6 126 òú ó ó ó ML2/13 108 òú ó ó ó ML4/13 158 òú ó ó ó ML1/14 85 òú ó ó ó ML9/9.5 52 òú ó ó ó ML9/10.5 53 òú ùòòòòòòò÷ ó ML8/19 37 òú ó ó ML2/12 107 òú ó ó ML3/8 128 òú ó ó ML7/8 3 òú ó ó ML7/11 6 òú ó ó ML7/6 1 òú ó ó ML4/12 157 ò÷ ó ó ML6/11 205 òø ó ó ML6/12 206 òú ó ó ML5/7 179 òú ó ó ML6/13 207 òú ó ó ML6/14 208 òú ó ó ML5/8 180 òú ó ó ML5/9 181 òú ó ó ML5/10 182 òú ó ó ML5/11 183 òú ó ó ML6/15 209 òú ó ó ML6/17 211 òú ó ó ML6/18 212 òú ó ó ML4/14 159 òú ó ó ML6/16 210 òôò÷ ó ML4/16 161 òú ó ML4/17 162 òú ó ML1/9 80 òú ó ML5/12 184 òú ó ML6/19 213 òú ó ML1/10 81 òú ó ML2/9 104 òú ó ML3/15 135 òú ó ML3/16 136 òú ó ML4/15 160 òú ó ML4/19 164 òú ó ML4/21 166 òú ó ML3/14 134 òú ó ML4/20 165 òú ó ML5/13 185 òú ó ML4/18 163 òú ó ML7/14 9 òú ó ML8/16 34 òú ó ML7/13 8 òú ó ML7/12 7 òú ó ML8/12 30 òú ó ML4/26 171 òú ó ML4/30 175 òú ó ML6/22 216 òú ó ML1/11 82 òú ó ML1/13 84 òú ó ML1/12 83 òú ó ML8/14 32 òú ó ML8/15 33 òú ó ML4/7 152 òú ó ML4/28 173 òú ó ML4/24 169 òú ó ML4/25 170 òú ó ML4/22 167 òú ó ML4/23 168 òú ó ML4/27 172 òú ó ML4/29 174 òú ó ML8/13 31 ò÷ ó ML3/22 142 òø ó ML3/23 143 òú ó ML3/21 141 òú ó ML3/24 144 òôòòòòòø ó ML3/17 137 ò÷ ó ó ML3/25 145 òø ó ó ML3/26 146 òú ó ó ML3/18 138 òú ùòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòò÷ ML3/20 140 òôòø ó ML3/19 139 ò÷ ùòø ó ML1/29 98 òûòú ó ó ML1/30 99 ò÷ ó ó ó ML1/26 95 òø ó ùò÷ ML1/27 96 òôò÷ ó ML1/28 97 ò÷ ó ML3/27 147 òòòòò÷

Appendix B7 – Statistical Results


Dendrogram distinguishing groundwater groups at Stuarts Point

Rescaled Distance Cluster Combine C A S E 0 5 10 15 20 25 Sample # +---------+---------+---------+---------+---------+ ML8/22 40 òø ML8/26 44 òú ML8/23 41 òú ML1/17 88 òú ML3/9 129 òú ML3/10 130 òú ML8/9 27 òú ML1/18 89 òú ML8/24 42 òú ML9/8.5 51 òú ML3/4 124 òú ML8/27 45 òú ML9/5.5 48 òú ML9/12.5 54 òú ML7/17 12 òú ML7/19 14 òú ML7/20 15 òú ML7/18 13 òú ML7/21 16 òú ML5/15 187 òú ML5/18 190 òú ML5/19 191 òú ML5/16 188 òú ML5/20 192 òú ML7/22 17 òú ML5/17 189 òú ML10/21 73 òú ML10/10 68 òú ML10/11 69 òú ML10/9 67 òú ML10/12 70 òú ML2/17 112 òôòòòòòø ML2/18 113 òú ó ML2/19 114 òú ó ML2/14 109 òú ó ML2/15 110 òú ó ML2/10 105 òú ó ML9/6.5 49 òú ó ML3/11 131 òú ó ML1/20 91 òú ó ML4/11 156 òú ó ML8/10 28 òú ó ML1/16 87 òú ó ML2/16 111 òú ó ML2/20 115 òú ó ML1/19 90 òú ó ML2/22 117 òú ó ML2/21 116 òú ó ML8/21 39 òú ó ML8/25 43 òú ó ML8/7 25 òú ó ML8/20 38 òú ó ML8/28 46 òú ó


C A S E 0 5 10 15 20 25 Sample # +---------+---------+---------+---------+---------+ ML1/21 92 òú ó ML2/23 118 òú ó ML1/15 86 òú ó ML4/8 153 òú ó ML4/9 154 òú ùòòòòòòòòòòòø ML4/10 155 òú ó ó ML2/24 119 òú ó ó ML3/5 125 òú ó ó ML2/25 120 òú ó ó ML2/26 121 òú ó ó ML2/27 122 òú ó ó ML2/11 106 ò÷ ó ó ML5/27 197 òø ó ó ML6/29 223 òú ó ó ML7/28 23 òú ó ó ML1/24 93 òú ó ó ML10/22 74 òú ó ó ML10/23 75 òú ó ó ML7/26 21 òú ó ó ML7/27 22 òú ó ó ML7/25 20 òú ó ó ML7/23 18 òôòø ó ó ML7/24 19 òú ó ó ó ML9/13.5 55 òú ó ó ó ML9/17.5 59 òú ó ó ó ML9/16.5 58 òú ó ó ó ML9/18.5 60 òú ó ó ó ML8/8 26 òú ó ó ó ML10/7 66 òú ó ó ó ML10/13 71 òú ó ó ó ML9/21.5 63 òú ó ó ó ML10/16 72 òú ùòòò÷ ó ML9/7.5 50 òú ó ó ML9/14.5 56 òú ó ó ML9/15.5 57 òú ó ó ML9/19.5 61 òú ó ó ML9/20.5 62 ò÷ ó ó ML5/30 200 òø ó ó ML6/30 224 òú ó ùòòòòòòòòòòòòòòòòòòòòòòòòòòòòòø ML5/28 198 òú ó ó ó ML5/29 199 òú ó ó ó ML10/6 65 òôò÷ ó ó ML1/25 94 òú ó ó ML10/5 64 ò÷ ó ó ML1/7 78 òø ó ó ML4/5 150 òú ó ó ML5/6 178 òú ó ó ML4/3 148 òú ó ó ML4/6 151 òú ó ó ML6/7 201 òú ó ó ML5/4 176 òú ó ó ML1/5 76 òú ó ó ML3/3 123 òôòòòòòòòòòø ó ó ML4/4 149 òú ó ó ó ML1/6 77 òú ó ó ó ML5/5 177 òú ó ó ó ML6/9 203 òú ó ó ó ML6/10 204 òú ó ó ó


C A S E 0 5 10 15 20 25 Sample # +---------+---------+---------+---------+---------+ ML2/5 100 òú ó ó ó ML2/6 101 òú ó ó ó ML2/8 103 òú ó ó ó ML6/8 202 òú ó ó ó ML2/7 102 ò÷ ó ó ó ML5/22 194 òø ó ó ó ML5/26 196 òú ó ó ó ML5/23 195 òú ó ó ó ML8/29 47 òú ó ó ó ML6/27 221 òú ó ó ó ML6/28 222 òú ó ó ó ML6/26 220 òú ó ó ó ML5/21 193 òú ó ó ó ML6/25 219 òú ó ó ó ML3/13 133 òú ó ó ó ML6/20 214 òú ùòòòòòòò÷ ó ML7/9 4 òú ó ó ML3/7 127 òôòø ó ó ML3/12 132 òú ó ó ó ML1/8 79 òú ó ó ó ML5/14 186 òú ó ó ó ML6/23 217 òú ó ó ó ML6/24 218 òú ó ó ó ML8/17 35 òú ó ó ó ML6/21 215 òú ó ó ó ML7/7 2 òú ó ó ó ML7/15 10 òú ó ó ó ML7/16 11 òú ó ó ó ML8/18 36 òú ó ó ó ML7/10 5 òú ó ó ó ML8/6 24 òú ó ó ó ML8/11 29 òú ó ó ó ML3/6 126 òú ó ó ó ML2/13 108 òú ó ó ó ML4/13 158 òú ó ó ó ML1/14 85 òú ó ó ó ML9/9.5 52 òú ó ó ó ML9/10.5 53 òú ùòòòòòòò÷ ó ML8/19 37 òú ó ó ML2/12 107 òú ó ó ML3/8 128 òú ó ó ML7/8 3 òú ó ó ML7/11 6 òú ó ó ML7/6 1 òú ó ó ML4/12 157 ò÷ ó ó ML6/11 205 òø ó ó ML6/12 206 òú ó ó ML5/7 179 òú ó ó ML6/13 207 òú ó ó ML6/14 208 òú ó ó ML5/8 180 òú ó ó ML5/9 181 òú ó ó ML5/10 182 òú ó ó ML5/11 183 òú ó ó ML6/15 209 òú ó ó ML6/17 211 òú ó ó ML6/18 212 òú ó ó


C A S E 0 5 10 15 20 25 Sample # +---------+---------+---------+---------+---------+ ML4/14 159 òú ó ó ML6/16 210 òôò÷ ó ML4/16 161 òú ó ML4/17 162 òú ó ML1/9 80 òú ó ML5/12 184 òú ó ML6/19 213 òú ó ML1/10 81 òú ó ML2/9 104 òú ó ML3/15 135 òú ó ML3/16 136 òú ó ML4/15 160 òú ó ML4/19 164 òú ó ML4/21 166 òú ó ML3/14 134 òú ó ML4/20 165 òú ó ML5/13 185 òú ó ML4/18 163 òú ó ML7/14 9 òú ó ML8/16 34 òú ó ML7/13 8 òú ó ML7/12 7 òú ó ML8/12 30 òú ó ML4/26 171 òú ó ML4/30 175 òú ó ML6/22 216 òú ó ML1/11 82 òú ó ML1/13 84 òú ó ML1/12 83 òú ó ML8/14 32 òú ó ML8/15 33 òú ó ML4/7 152 òú ó ML4/28 173 òú ó ML4/24 169 òú ó ML4/25 170 òú ó ML4/22 167 òú ó ML4/23 168 òú ó ML4/27 172 òú ó ML4/29 174 òú ó ML8/13 31 ò÷ ó ML3/22 142 òø ó ML3/23 143 òú ó ML3/21 141 òú ó ML3/24 144 òôòòòòòø ó ML3/17 137 ò÷ ó ó ML3/25 145 òø ó ó ML3/26 146 òú ó ó ML3/18 138 òú ùòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòò÷ ML3/20 140 òôòø ó ML3/19 139 ò÷ ùòø ó ML1/29 98 òûòú ó ó ML1/30 99 ò÷ ó ó ó ML1/26 95 òø ó ùò÷ ML1/27 96 òôò÷ ó ML1/28 97 ò÷ ó ML3/27 147 òòòòò÷



Dendrogram showing element similarities for the Shallow Groundwaters

Rescaled Distance Cluster Combine 0 5 10 15 20 25

Element # +---------+---------+---------+---------+---------+ Ba 4 òûòòòòòø Ga 12 ò÷ ùòø Ng 34 òòòûòø ó ó NO3 42 òòò÷ ùò÷ ó Rb 24 òòòø ó ùòòòòòòòòòòòòòòòòòòòòòø K 36 òòòôò÷ ó ó SO4 39 òòò÷ ó ó Ag 1 òòòòòòòòò÷ ó P 21 òòòûòø ó EC 43 òòò÷ ùòòòø ùòòòòòòòòòòòòòòòòòø Temp 44 òòòòò÷ ùòòòòòòòòòòòòòòòø ó ó Mo 18 òòòòòòòûò÷ ó ó ó Pb 22 òòòòòòò÷ ó ó ó Cr 7 òòòø ó ó ó V 29 òòòôòòòø ùòòòòò÷ ó Zr 32 òòò÷ ùòòòòòòòòòòòø ó ó Al 2 òòòòòûò÷ ó ó ó U 28 òòòòò÷ ó ó ó Na 35 òòòûòòòòòòòø ó ó ó Cl 38 òòò÷ ùòòòø ùòòòòò÷ ó Li 16 òòòòòòòûòòòú ó ó ó pH 45 òòòòòòò÷ ó ó ó ó Ge 14 òòòûòòòø ó ó ó ó Mn 17 òòò÷ ùòòò÷ ùòòò÷ ó Co 6 òòòòòûò÷ ó ó Ni 20 òòòòò÷ ó ó Sr 27 òûòòòòòø ó ó Ca 33 ò÷ ùòòòòòòòú ó HCO3 37 òòòòòòò÷ ó ó As 3 òòòûòòòòòø ó ó Fe 40 òòò÷ ùòòòòò÷ ó Cs 8 òòòòòûòø ó ó Sc 25 òòòòò÷ ùò÷ ó NH4 41 òòòòòòò÷ ó Dy 9 òø ó Y 30 òôòòòø ó Er 10 òú ó ó Yb 31 ò÷ ùòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòò÷ Ce 5 òûòø ó La 15 ò÷ ó ó Nd 19 òø ùò÷ Pr 23 òú ó Gd 13 òôò÷ Sm 26 òú Eu 11 ò÷

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0.25

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40.

10-0

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Al

1.00

0.13

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90.

620.

200.

480.

400.

500.

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0.55

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0.53

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130.

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0.53

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20.

220.

19-0

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0.14

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80.

110.

28-0

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0.24

0.04

0.54

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0.21

0.31

0.23

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80.

290.

46-0

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0.40

0.86

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0.00

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90.

100.

36B

a1.

000.

310.

28-0

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0.66

0.28

0.32

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0.33

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60.

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000.

330.

900.

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290.

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920.

340.

360.

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40.

270.

380.

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600.

460.

46-0

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0.58

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56-0

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Cr

1.00

0.18

0.23

0.20

0.17

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21-0

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0.10

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000.

270.

16-0

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0.15

0.11

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80.

450.

210.

180.

610.

870.

210.

190.

860.

14-0

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80.

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370.

350.

490.

680.

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0.74

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340.

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280.

990.

780.

960.

230.

64-0

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0.99

0.53

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80.

120.

980.

350.

730.

990.

560.

45-0

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0.99

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0.44

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960.

320.

970.

770.

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220.

66-0

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0.97

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090.

960.

410.

750.

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580.

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0.33

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180.

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450.

990.

800.

960.

240.

66-0

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0.99

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30.

160.

980.

490.

750.

990.

580.

32-0

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0.96

0.94

0.28

0.47

0.64

0.26

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20.

350.

580.

010.

130.

64-0

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60.

03G

a1.

000.

330.

290.

350.

210.

47-0

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0.34

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6-0

.36

0.00

0.33

0.90

0.47

0.34

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0.29

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90.

270.

60-0

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0.87

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70.

080.

720.

240.

160.

72-0

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d1.

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970.

240.

64-0

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1.00

0.55

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400.

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41-0

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0.98

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0.37

0.43

0.59

0.31

0.34

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0.37

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50.

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80.

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790.

340.

770.

150.

810.

64-0

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0.20

0.80

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0.53

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740.

110.

540.

680.

230.

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56-0

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65-0

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370.

63-0

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0.98

0.57

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80.

140.

990.

420.

710.

970.

460.

42-0

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0.96

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0.29

0.34

0.62

0.26

0.32

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60.

350.

46-0

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0.05

0.61

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000.

37-0

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0.25

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0.27

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0.31

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0.26

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0.36

0.05

0.06

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70.

170.

03-0

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0.01

0.32

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0.38

Mn

1.00

0.04

0.63

0.41

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20.

180.

610.

510.

700.

630.

590.

12-0

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0.68

0.66

0.15

0.57

0.67

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0.46

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0.22

0.70

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0.35

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50.

400.

09-0

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0.26

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0.53

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0.14

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950.

330.

420.

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280.

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0.55

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570.

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260.

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220.

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0.15

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120.

120.

070.

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0.08

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0.11

0.17

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50.

030.

320.

190.

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1.00

0.40

0.70

0.99

0.52

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50.

970.

940.

340.

400.

580.

260.

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39-0

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80.

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0.69

0.23

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60.

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790.

250.

310.

82-0

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0.73

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0.76

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0.98

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90.

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1.00

0.01

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10.

560.

550.

240.

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400.

230.

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0.92

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mp

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1.00

Alth

ough

cor

rela

tions

bet

wee

n ar

seni

c an

d ot

her e

lem

ents

are

hig

hlig

hted

her

e, o

ther

sig

nific

ant c

orre

latio

ns e

xist

bet

wee

en o

ther

ele

men

ts in

thes

e w

ater

s. T

hese

oth

er c

orre

latio

ns a

re c

urre

ntly

the

topi

c of

furt

her s

tudy

by

the

auth

or h

erei

n.

App

endi

x B

7 - S

tatis

tical

Res

ults



Dendrogram showing element similarities for the Barrier Sand Groundwaters

Rescaled Distance Cluster Combine 0 5 10 15 20 25

Element # +---------+---------+---------+---------+---------+ Nd 19 òø Sm 26 òú Gd 13 òôòòòø Ce 5 òú ó La 15 òú ó Pr 23 ò÷ ùòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòø Dy 9 òø ó ó Y 30 òôòø ó ó Er 10 ò÷ ó ó ó Yb 31 òòòôò÷ ó Eu 11 òòò÷ ó Pb 22 òòòòòûòòòø ó Temp 44 òòòòò÷ ó ó Ge 14 òòòø ùòòòòòòòòòòòòòòòòòòòòòø ó Mo 18 òòòú ó ó ó NO3 42 òòòôòòòø ó ó ó Ag 1 òòò÷ ùò÷ ó ó As 3 òòòòòûò÷ ó ó pH 45 òòòòò÷ ó ó Sr 27 òø ó ó Ca 33 òôòø ùòòòòòòòòòòòòòòòòò÷ HCO3 37 ò÷ ùòòòòòø ó Mn 17 òòòú ó ó Fe 40 òòò÷ ùòòòòòòòòòòòø ó Ba 4 òòòûòø ó ó ó Ga 12 òòò÷ ùòòò÷ ó ó Ni 20 òòòûò÷ ó ó Mg 34 òòò÷ ó ó Co 6 òòòûòø ùòòòòòòòòò÷ Sc 25 òòò÷ ùòø ó U 28 òòòòò÷ ùòòòø ó Li 16 òòòòòûò÷ ó ó P 21 òòòòò÷ ó ó Cs 8 òòòòòûòø ùòòòòòòòòò÷ NH4 41 òòòòò÷ ùòòòú Zr 32 òòòòòòò÷ ó Cr 7 òòòûòø ó V 29 òòò÷ ùòø ó Al 2 òòòòò÷ ùòø ó SO4 39 òòòòòòòú ó ó EC 43 òòòòòòò÷ ùò÷ Rb 24 òûòòòòòø ó K 36 ò÷ ùò÷ Na 35 òòòòòûò÷ Cl 38 òòòòò÷

Pear

sons

Cor

rela

tions

bet

wee

n el

emen

ts in

the

Bar

rier S

and

Gro

undw

ater

s

AgAl

AsB

aC

eC

oC

rC

sD

yEr

EuG

aG

dG

eLa

LiM

nM

oN

dN

iP

PbPr

Rb

ScSm

SrU

VY

YbZr

Ca

Mg

Na

KH

CO

3C

lSO

4Fe

NH

4N

O3

ECTe

mp

pHAg

1.00

-0.0

80.

39-0

.39

-0.3

7-0

.48

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80.

27-0

.30

-0.3

0-0

.31

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2-0

.36

0.70

-0.2

70.

160.

150.

58-0

.30

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8-0

.01

0.29

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5-0

.58

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5-0

.35

0.07

-0.2

8-0

.51

-0.3

4-0

.43

0.29

0.14

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7-0

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50.

28-0

.45

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1-0

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60.

69-0

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0.07

0.31

Al1.

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10.

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130.

590.

140.

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520.

35-0

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0.55

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60.

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380.

490.

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29-0

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0.08

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80.

120.

15-0

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0.05

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18-0

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2As

1.00

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6-0

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9-0

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0.16

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2-0

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80.

47-0

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20.

040.

62-0

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0.09

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60.

130.

03-0

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2-0

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0.05

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4-0

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9-0

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0.45

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60.

250.

35B

a1.

00-0

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0.33

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0.24

0.82

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2-0

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40.

400.

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60.

610.

09-0

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90.

720.

37-0

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0.65

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50.

310.

110.

14-0

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0.55

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0.08

0.72

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60.

26-0

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e1.

000.

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71-0

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0.91

0.86

0.75

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20.

96-0

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0.97

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0-0

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140.

000.

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40.

33-0

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Co

1.00

0.36

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00.

160.

190.

210.

330.

19-0

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0.11

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0.13

0.58

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10.

090.

510.

650.

190.

000.

590.

630.

160.

39-0

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0.70

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70.

360.

320.

200.

23-0

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0.22

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5C

r1.

00-0

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0.68

0.68

0.49

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10.

70-0

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0.63

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80.

660.

12-0

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0.18

0.61

0.28

0.17

0.69

-0.3

50.

350.

670.

680.

74-0

.02

-0.3

50.

090.

240.

42-0

.44

0.27

0.36

-0.0

80.

01-0

.44

0.25

0.05

-0.2

0C

s1.

00-0

.13

-0.1

6-0

.07

0.36

-0.1

70.

07-0

.09

0.17

0.41

0.01

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60.

04-0

.05

-0.1

7-0

.13

0.25

-0.1

4-0

.20

0.47

-0.4

7-0

.16

-0.1

0-0

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0.31

0.40

0.21

0.06

-0.0

10.

400.

28-0

.29

0.30

0.34

0.01

-0.2

4-0

.26

-0.0

3D

y1.

000.

940.

86-0

.07

0.96

-0.2

70.

91-0

.30

-0.3

2-0

.34

0.95

-0.0

7-0

.31

0.04

0.91

0.11

0.05

0.94

-0.3

60.

280.

560.

970.

870.

20-0

.39

-0.0

20.

080.

20-0

.37

0.05

0.11

-0.0

90.

08-0

.32

0.37

-0.1

20.

12Er

1.00

0.74

-0.0

60.

93-0

.28

0.85

-0.2

7-0

.30

-0.4

20.

890.

02-0

.27

0.06

0.84

0.11

0.12

0.90

-0.3

00.

230.

620.

950.

880.

22-0

.33

0.05

0.06

0.18

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90.

070.

04-0

.03

0.10

-0.3

50.

33-0

.13

-0.0

2Eu

1.00

0.22

0.82

-0.2

80.

80-0

.09

-0.1

8-0

.28

0.82

0.04

-0.2

3-0

.11

0.80

0.20

0.15

0.82

-0.2

20.

310.

430.

830.

740.

08-0

.25

0.07

0.11

0.28

-0.2

50.

090.

150.

020.

12-0

.31

0.39

-0.1

80.

20G

a1.

00-0

.13

-0.0

6-0

.13

0.59

0.54

-0.1

4-0

.16

0.56

0.24

-0.2

0-0

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0.42

0.56

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40.

560.

070.

20-0

.01

0.07

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70.

520.

550.

130.

330.

560.

22-0

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0.56

0.35

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00.

13-0

.31

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9G

d1.

00-0

.29

0.96

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5-0

.38

-0.3

40.

98-0

.04

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20.

030.

940.

120.

030.

98-0

.38

0.23

0.54

0.95

0.86

0.13

-0.4

0-0

.04

0.03

0.23

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20.

050.

16-0

.14

0.05

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20.

36-0

.10

0.12

Ge

1.00

-0.2

50.

07-0

.09

0.80

-0.2

5-0

.33

-0.0

10.

32-0

.21

-0.6

8-0

.10

-0.3

2-0

.14

-0.1

2-0

.52

-0.3

3-0

.38

0.20

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6-0

.11

-0.1

1-0

.53

0.11

-0.4

5-0

.28

-0.3

8-0

.20

0.78

-0.2

60.

280.

41La

1.00

-0.3

5-0

.42

-0.3

10.

98-0

.15

-0.3

00.

010.

970.

11-0

.04

0.96

-0.4

10.

160.

440.

880.

770.

09-0

.42

-0.0

7-0

.05

0.20

-0.4

5-0

.03

0.20

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90.

04-0

.27

0.35

-0.1

90.

17Li

1.00

0.54

0.06

-0.3

90.

380.

46-0

.05

-0.4

00.

130.

62-0

.36

0.48

0.02

0.11

-0.2

9-0

.10

-0.0

70.

490.

240.

02-0

.08

0.60

-0.0

4-0

.14

0.47

0.20

0.04

-0.1

6-0

.17

-0.0

5M

n1.

00-0

.01

-0.4

20.

47-0

.08

-0.2

7-0

.42

0.31

0.22

-0.4

00.

84-0

.31

-0.0

4-0

.26

-0.2

60.

090.

820.

35-0

.20

0.09

0.76

0.23

-0.3

60.

760.

31-0

.07

-0.0

6-0

.08

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3M

o1.

00-0

.30

-0.3

4-0

.08

0.28

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5-0

.64

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2-0

.35

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4-0

.07

-0.5

7-0

.40

-0.5

00.

19-0

.07

-0.2

9-0

.13

-0.4

60.

04-0

.37

-0.2

1-0

.43

-0.1

90.

75-0

.40

0.39

0.45

Nd

1.00

-0.1

5-0

.33

0.05

0.98

0.09

-0.0

40.

98-0

.44

0.21

0.46

0.92

0.80

0.11

-0.4

5-0

.09

-0.0

20.

20-0

.46

-0.0

30.

18-0

.20

0.02

-0.2

60.

33-0

.10

0.17

Ni

1.00

0.02

-0.0

6-0

.21

0.39

0.58

-0.0

80.

590.

180.

360.

010.

21-0

.24

0.62

0.62

0.07

0.38

0.49

0.43

0.00

0.64

0.26

-0.4

40.

12-0

.01

-0.4

2P

1.00

-0.0

4-0

.35

0.15

0.49

-0.3

0-0

.10

0.26

0.16

-0.3

4-0

.11

-0.1

2-0

.11

-0.0

40.

160.

05-0

.03

-0.1

30.

22-0

.04

0.03

-0.0

2-0

.03

-0.2

4-0

.02

Pb1.

000.

05-0

.47

-0.0

50.

04-0

.30

0.10

-0.1

20.

060.

010.

13-0

.25

-0.1

90.

15-0

.32

-0.0

6-0

.20

-0.2

2-0

.30

-0.2

70.

35-0

.11

0.32

0.12

Pr1.

000.

06-0

.08

0.95

-0.4

40.

190.

400.

880.

740.

10-0

.46

-0.1

0-0

.04

0.17

-0.4

6-0

.07

0.17

-0.2

30.

01-0

.22

0.33

-0.0

80.

20R

b1.

000.

330.

140.

330.

070.

540.

170.

25-0

.09

0.22

0.45

0.22

0.85

0.06

0.58

0.28

0.51

0.30

-0.7

60.

24-0

.30

-0.3

6Sc

1.00

0.01

0.20

0.49

0.61

0.03

0.32

-0.1

80.

240.

510.

100.

330.

260.

080.

210.

380.

16-0

.25

0.14

-0.0

2-0

.12

Sm1.

00-0

.43

0.25

0.52

0.93

0.83

0.10

-0.4

4-0

.09

0.02

0.25

-0.4

60.

020.

19-0

.15

0.04

-0.3

20.

33-0

.08

0.11

Sr1.

00-0

.44

-0.0

7-0

.27

-0.2

70.

070.

960.

49-0

.02

0.09

0.88

0.36

-0.2

70.

760.

27-0

.17

-0.0

6-0

.19

-0.3

1U

1.00

0.43

0.21

0.40

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9-0

.36

0.14

0.24

0.31

-0.3

2-0

.05

0.35

-0.1

70.

04-0

.27

0.15

0.25

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3V

1.00

0.55

0.75

0.02

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70.

360.

110.

56-0

.14

0.28

0.30

0.26

0.29

-0.6

50.

31-0

.11

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8Y

1.00

0.87

0.22

-0.3

20.

000.

110.

24-0

.30

0.12

0.05

-0.0

10.

09-0

.38

0.39

-0.1

30.

03Yb

1.00

0.04

-0.2

70.

150.

110.

33-0

.30

0.12

0.26

0.05

0.23

-0.4

80.

37-0

.12

-0.0

9Zr

1.00

0.02

-0.1

60.

20-0

.21

0.15

0.11

-0.5

4-0

.06

0.05

0.04

-0.0

1-0

.07

0.19

Ca

1.00

0.51

-0.1

10.

020.

900.

26-0

.25

0.74

0.25

-0.0

9-0

.15

-0.0

9-0

.32

Mg

1.00

0.08

0.52

0.48

0.32

0.09

0.51

0.12

-0.3

20.

090.

00-0

.43

Na

1.00

0.23

-0.0

20.

570.

07-0

.05

0.01

-0.3

00.

04-0

.03

-0.1

0K

1.00

-0.1

50.

510.

450.

260.

16-0

.69

0.32

-0.0

9-0

.30

HC

O3

1.00

0.09

-0.4

60.

650.

180.

07-0

.20

-0.1

0-0

.25

Cl

1.00

-0.0

40.

390.

32-0

.59

0.06

-0.0

8-0

.32

SO4

1.00

-0.2

0-0

.08

-0.2

00.

15-0

.08

-0.0

3Fe

1.00

0.36

-0.4

10.

11-0

.21

-0.3

5N

H4

1.00

-0.3

70.

05-0

.22

-0.1

3N

O3

1.00

-0.4

20.

220.

41EC

1.00

-0.3

10.

22Te

mp

1.00

-0.0

5pH

1.00

Alth

ough

cor

rela

tions

bet

wee

n ar

seni

c an

d ot

her e

lem

ents

are

hig

hlig

hted

her

e, o

ther

sig

nific

ant c

orre

latio

ns e

xist

bet

wee

en o

ther

ele

men

ts in

thes

e w

ater

s. T

hese

oth

er c

orre

latio

ns a

re c

urre

ntly

the

topi

c of

furt

her s

tudy

by

the

auth

or h

erei

n.

App

endi

x B

7 - S

tatis

tical

Res

ults


* * * * * * H I E R A R C H I C A L C L U S T E R A N A L Y S I S * * * * *

Dendrogram showing element similarities for the Fluvial Sand Groundwaters

Rescaled Distance Cluster Combine

0 5 10 15 20 25 Element # +---------+---------+---------+---------+---------+ Ba 4 òûòòòòòòòòòòòòòø Ga 12 ò÷ ùòòòòòø Rb 24 òòòûòòòòòø ó ó K 36 òòò÷ ùòòòòò÷ ó Co 6 òòòòòûòø ó ó V 29 òòòòò÷ ùò÷ ùòòòòòòòòòòòø Al 2 òòòòòòò÷ ó ó Sr 27 òûòòòø ó ó Ca 33 ò÷ ùòòòòòòòø ó ó Ni 20 òòòòò÷ ó ó ó Li 16 òòòûòø ùòòòòòòò÷ ó U 28 òòò÷ ùòòòø ó ó Mg 34 òòòòò÷ ùòòò÷ ó SO4 39 òòòòòûòø ó ó NO3 42 òòòòò÷ ùò÷ ùòòòòòòòòòòòòòòòø Cl 38 òòòòòòò÷ ó ó As 3 òòòòòø ó ó Ge 14 òòòòòôòø ó ó Mo 18 òòòòò÷ ùòòòø ó ó HCO3 37 òòòòòòò÷ ùòø ó ó Sc 25 òòòòòûòòòø ó ó ó ó Na 35 òòòòò÷ ùò÷ ó ó ó Cr 7 òòòòòòòûò÷ ó ó ó Temp 44 òòòòòòò÷ ùòòòòòòòòòòòòòòòòòòò÷ ó Ag 1 òòòòòòòø ó ó Zr 32 òòòòòòòú ó ó Cs 8 òòòòòòòôòòòòòú ó Pb 22 òòòòòòò÷ ó ó EC 43 òòòòòûòòòòòø ó ó pH 45 òòòòò÷ ùò÷ ó Mn 17 òòòòòûòòòø ó ó Fe 40 òòòòò÷ ùò÷ ó P 21 òòòòòòòûò÷ ó NH4 41 òòòòòòò÷ ó Ce 5 òòòûòòòø ó La 15 òòò÷ ó ó Er 10 òòòø ùòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòò÷ Eu 11 òòòôòø ó Yb 31 òòò÷ ó ó Pr 23 òø ùò÷ Sm 26 òôòø ó Nd 19 ò÷ ùò÷ Dy 9 òûòú Y 30 ò÷ ó Gd 13 òòò÷

Pear

sons

Cor

rela

tions

bet

wee

n el

emen

ts in

the

Fluv

ial S

and

Gro

undw

ater

s

Ag

Al

As

Ba

Ce

Co

Cr

Cs

Dy

ErEu

Ga

Gd

Ge

LaLi

Mn

Mo

Nd

Ni

PPb

PrR

bSc

SmSr

UV

YYb

ZrC

aM

gN

aK

HC

O3

Cl

SO4

FeN

H4

NO

3EC

Tem

ppH

Ag

1.00

-0.2

90.

160.

19-0

.56

-0.6

1-0

.33

0.42

-0.0

6-0

.01

-0.0

70.

30-0

.16

-0.1

0-0

.25

0.01

0.37

0.05

-0.1

6-0

.03

-0.1

40.

35-0

.17

0.09

-0.1

9-0

.18

0.23

-0.2

0-0

.21

0.10

-0.2

00.

440.

180.

030.

19-0

.28

0.22

0.17

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70.

04-0

.21

0.24

0.15

0.09

0.32

Al

1.00

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00.

040.

560.

370.

240.

110.

380.

260.

180.

050.

390.

060.

51-0

.25

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1-0

.40

0.42

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2-0

.17

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30.

390.

10-0

.04

0.39

-0.2

5-0

.06

0.48

0.25

0.28

0.00

-0.2

1-0

.04

-0.3

40.

12-0

.31

-0.1

10.

23-0

.17

0.21

-0.0

4-0

.07

-0.1

90.

10A

s1.

000.

03-0

.41

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8-0

.17

0.15

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0-0

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70.

09-0

.16

0.65

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5-0

.35

0.21

0.58

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5-0

.16

-0.0

20.

04-0

.30

-0.4

80.

18-0

.26

-0.0

5-0

.11

-0.6

4-0

.12

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90.

230.

10-0

.33

0.33

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80.

53-0

.30

-0.5

50.

050.

23-0

.53

-0.0

80.

060.

31B

a1.

000.

060.

03-0

.08

0.39

0.35

0.37

0.52

0.97

0.20

0.23

0.09

-0.1

9-0

.09

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30.

20-0

.09

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10.

140.

200.

520.

290.

27-0

.27

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2-0

.02

0.46

0.38

0.12

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70.

020.

210.

34-0

.14

-0.0

3-0

.32

0.31

-0.0

50.

040.

11-0

.14

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1C

e1.

000.

650.

31-0

.10

0.67

0.67

0.66

-0.0

20.

730.

030.

850.

15-0

.24

-0.3

20.

780.

060.

00-0

.28

0.82

0.34

0.09

0.80

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50.

110.

380.

560.

72-0

.35

-0.3

50.

14-0

.21

0.47

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70.

040.

35-0

.05

-0.0

10.

130.

02-0

.07

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3C

o1.

000.

33-0

.14

0.30

0.34

0.40

-0.0

80.

360.

050.

420.

29-0

.26

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90.

360.

41-0

.10

-0.3

00.

370.

340.

350.

390.

030.

410.

500.

200.

48-0

.36

0.04

0.27

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80.

61-0

.36

0.11

0.40

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3-0

.01

0.20

0.01

0.05

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4C

r1.

00-0

.08

0.06

0.16

0.11

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00.

190.

140.

08-0

.04

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8-0

.04

0.12

0.17

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00.

180.

110.

050.

220.

16-0

.08

0.00

0.36

0.15

0.23

0.03

-0.0

2-0

.01

0.14

0.18

0.05

0.09

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5-0

.04

0.01

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6-0

.02

0.24

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3C

s1.

000.

320.

340.

270.

480.

230.

330.

090.

09-0

.02

0.23

0.17

0.15

-0.0

20.

220.

160.

360.

330.

160.

01-0

.03

0.05

0.40

0.18

0.44

0.07

0.21

0.26

0.18

0.12

0.07

-0.2

3-0

.02

-0.0

40.

10-0

.02

0.24

0.16

Dy

1.00

0.84

0.76

0.36

0.89

0.16

0.71

0.02

0.09

-0.1

70.

82-0

.01

-0.1

0-0

.09

0.84

0.40

0.11

0.85

-0.3

8-0

.12

0.17

0.89

0.79

0.11

-0.3

2-0

.06

0.00

0.24

-0.5

4-0

.03

0.11

0.07

0.00

0.14

0.03

0.14

-0.0

3Er

1.00

0.85

0.35

0.82

0.11

0.71

0.17

0.03

-0.1

50.

810.

08-0

.10

-0.0

70.

840.

460.

190.

84-0

.37

0.03

0.17

0.86

0.82

0.08

-0.3

30.

100.

060.

36-0

.57

0.06

0.10

0.05

-0.1

60.

200.

100.

16-0

.03

Eu1.

000.

470.

730.

190.

670.

220.

05-0

.07

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210.

240.

05-0

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0.47

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500.

330.

26-0

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230.

23-0

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32-0

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340.

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230.

140.

410.

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310.

19-0

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80.

280.

180.

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70.

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CO

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mp

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3pH

1.00

Alth

ough

cor

rela

tions

bet

wee

n ar

seni

c an

d ot

her e

lem

ents

are

hig

hlig

hted

her

e, o

ther

sig

nific

ant c

orre

latio

ns e

xist

bet

wee

en o

ther

ele

men

ts in

thes

e w

ater

s. T

hese

oth

er c

orre

latio

ns a

re c

urre

ntly

the

topi

c of

furt

her s

tudy

by

the

auth

or h

erei

n.

App

endi

x B

7 - S

tatis

tical

Res

ults



Dendrogram showing element similarities for the Fluvial Sand & Clay Groundwaters

Rescaled Distance Cluster Combine 0 5 10 15 20 25 Element # +---------+---------+---------+---------+---------+ Na 35 òø Cl 38 òôòòòòòø K 36 ò÷ ùòòòø Cr 7 òûòòòø ó ó Li 16 ò÷ ùò÷ ó P 21 òûòø ó ùòòòø Sc 25 ò÷ ùò÷ ó ó V 29 òòò÷ ó ó U 28 òòòûòòòòòòò÷ ó pH 45 òòò÷ ùòòòòòòòòòòòòòø Al 2 òòòûòòòòòø ó ó Cs 8 òòò÷ ó ó ó Ba 4 òûòø ó ó ó Ga 12 ò÷ ó ùòòòòò÷ ó Rb 24 òòòôòø ó ó Eu 11 òòò÷ ùòø ó ó EC 43 òòòòò÷ ùò÷ ó NH4 41 òòòòòòò÷ ùòòòòòòòòòòòòòòòòòòòø Mn 17 òòòø ó ó HCO3 37 òòòú ó ó Mg 34 òòòú ó ó Sr 27 òø ùòø ó ó Ca 33 òôòú ó ó ó Ni 20 ò÷ ó ùòòòòòòòòòòòòòø ó ó Co 6 òòò÷ ó ó ó ó SO4 39 òòòòò÷ ó ó ó Ag 1 òòòòòø ùòòòòòòòòò÷ ó Pb 22 òòòòòú ó ó Fe 40 òòòòòôòòòø ó ó As 3 òòòòò÷ ó ó ó Mo 18 òòòø ùòòòòòòòòò÷ ó Temp 44 òòòôòòòø ó ó Zr 32 òòò÷ ùò÷ ó Ge 14 òòòòòûò÷ ó NO3 42 òòòòò÷ ó Ce 5 òûòòòø ó La 15 ò÷ ó ó Er 10 òûòø ùòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòòò÷ Yb 31 ò÷ ó ó Nd 19 òø ùò÷ Pr 23 òú ó Sm 26 òôò÷ Y 30 òú Gd 13 òú Dy 9 ò÷

Pear

sons

Cor

rela

tions

bet

wee

n el

emen

ts in

the

Fluv

ial S

and

& C

lay

Gro

undw

ater

s

Ag

Al

As

Ba

Ce

Co

Cr

Cs

Dy

ErEu

Ga

Gd

Ge

LaLi

Mn

Mo

Nd

Ni

PPb

PrR

bSc

SmSr

UV

YYb

ZrC

aM

gN

aK

HC

O3

Cl

SO4

FeN

H4

NO

3EC

Tem

ppH

Ag

1.00

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46-0

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60.

21-0

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0.02

0.15

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20.

120.

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0.08

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70.

280.

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1-0

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60.

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380.

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Al

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10.

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50.

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40.

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170.

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20-0

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10.

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210.

100.

250.

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14-0

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Ba

1.00

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40.

420.

330.

44-0

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710.

93-0

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30.

380.

150.

56-0

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0.66

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550.

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1.00

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0.85

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85-0

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81-0

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0.80

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0.11

0.54

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21-0

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0.16

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60.

750.

280.

32-0

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10.

230.

770.

550.

200.

250.

680.

350.

280.

320.

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070.

500.

27-0

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Cr

1.00

0.15

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0.16

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30.

84-0

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0.68

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0.30

0.67

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40.

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290.

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90.

450.

720.

660.

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260.

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0.20

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0.52

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0.11

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0.06

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0.34

0.00

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60.

19-0

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50.

01D

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000.

840.

03-0

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0.91

0.23

0.83

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9-0

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0.25

0.89

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10.

86-0

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40.

87-0

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0.92

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0.87

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84-0

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50.

86-0

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0.90

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0.28

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0.10

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05-0

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Eu1.

000.

760.

020.

08-0

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0.17

0.15

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0-0

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0.24

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40.

020.

560.

410.

050.

080.

170.

510.

01-0

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0.26

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0.39

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0.28

0.49

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0.70

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10.

370.

270.

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00.

230.

280.

600.

520.

600.

450.

600.

160.

280.

33-0

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0.71

0.25

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d1.

000.

220.

77-0

.40

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20.

300.

91-0

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4-0

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0.88

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0.93

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40.

910.

830.

30-0

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80.

760.

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6-0

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3-0

.08

Li1.

00-0

.01

-0.8

1-0

.38

0.10

0.78

-0.3

1-0

.43

0.38

0.76

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10.

050.

260.

69-0

.53

-0.5

4-0

.64

-0.1

60.

520.

880.

800.

160.

83-0

.11

0.09

0.37

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0-0

.22

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00.

16M

n1.

000.

38-0

.11

0.78

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40.

02-0

.11

0.46

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5-0

.07

0.86

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50.

11-0

.01

0.01

0.35

0.83

0.69

0.21

0.31

0.79

0.36

0.60

0.41

0.17

0.18

0.50

0.40

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6M

o1.

000.

240.

15-0

.72

0.31

0.26

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4-0

.62

0.33

0.25

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0-0

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0.42

0.43

0.65

0.39

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9-0

.56

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00.

14-0

.47

0.24

0.12

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80.

400.

410.

69-0

.23

Nd

1.00

-0.2

0-0

.39

-0.0

20.

93-0

.15

-0.5

60.

90-0

.08

-0.3

7-0

.28

0.90

0.85

0.24

-0.0

6-0

.15

-0.2

9-0

.30

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4-0

.30

0.34

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2-0

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0.10

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0-0

.03

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5N

i1.

00-0

.28

-0.0

4-0

.21

0.48

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1-0

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0.94

0.24

0.26

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4-0

.09

0.20

0.94

0.72

0.22

0.36

0.80

0.36

0.65

0.17

0.20

0.17

0.40

0.18

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9P

1.00

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3-0

.43

0.11

0.87

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1-0

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0.44

0.69

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1-0

.50

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2-0

.51

0.11

0.67

0.51

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90.

57-0

.50

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70.

20-0

.35

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7-0

.62

0.41

Pb1.

000.

01-0

.13

-0.1

7-0

.01

-0.0

6-0

.09

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60.

060.

040.

170.

03-0

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5-0

.41

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4-0

.29

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00.

21-0

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0.07

0.10

0.24

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0Pr

1.00

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6-0

.56

0.91

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1-0

.37

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80.

920.

880.

30-0

.10

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7-0

.33

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1-0

.17

-0.3

30.

340.

00-0

.14

0.10

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60.

05-0

.26

Rb

1.00

0.34

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90.

490.

070.

48-0

.11

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8-0

.03

0.36

0.66

0.60

0.71

0.62

0.65

0.41

0.10

0.36

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80.

50-0

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2Sc

1.00

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1-0

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0.61

0.81

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3-0

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0-0

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0.31

0.69

0.61

0.07

0.64

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40.

150.

34-0

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0.16

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1.00

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1-0

.29

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60.

930.

860.

34-0

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5-0

.31

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9-0

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20.

330.

01-0

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0.09

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00.

05-0

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000.

030.

17-0

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0.00

0.20

0.96

0.76

0.26

0.37

0.85

0.40

0.73

0.19

0.19

0.20

0.37

0.25

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1.00

0.62

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6-0

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0.02

0.06

0.15

0.23

0.23

0.20

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31-0

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0.58

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1-0

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0.03

0.60

0.78

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0.39

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48-0

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90.

22Y

1.00

0.90

0.39

0.01

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4-0

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60.

40-0

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30.

11-0

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0.17

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1.00

0.44

0.05

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70.

37-0

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00.

23-0

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0.15

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8Zr

1.00

0.31

0.00

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3-0

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0.25

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40.

160.

20-0

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0.45

0.52

0.62

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7C

a1.

000.

590.

020.

140.

770.

170.

690.

180.

080.

240.

410.

34-0

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Mg

1.00

0.73

0.81

0.76

0.81

0.55

0.26

0.45

0.07

0.36

0.00

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6N

a1.

000.

930.

390.

970.

100.

140.

46-0

.28

0.07

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80.

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1.00

0.45

0.94

0.28

0.15

0.44

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70.

21-0

.31

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CO

31.

000.

470.

500.

110.

340.

090.

440.

26-0

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Cl

1.00

0.21

0.26

0.43

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10.

17-0

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3SO

41.

000.

040.

080.

150.

140.

14-0

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Fe1.

000.

05-0

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0.18

0.07

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3N

H4

1.00

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30.

190.

000.

02N

O3

1.00

0.23

0.30

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7EC

1.00

0.53

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7Te

mp

1.00

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7pH

1.00

Alth

ough

cor

rela

tions

bet

wee

n ar

seni

c an

d ot

her e

lem

ents

are

hig

hlig

hted

her

e, o

ther

sig

nific

ant c

orre

latio

ns e

xist

bet

wee

en o

ther

ele

men

ts in

thes

e w

ater

s. T

hese

oth

er c

orre

latio

ns a

re c

urre

ntly

the

topi

c of

furt

her s

tudy

by

the

auth

or h

erei

n.

App

endi

x B

7 - S

tatis

tical

Res

ults



Dendrogram showing element similarities for the Seawater Intrusion Groundwaters

Rescaled Distance Cluster Combine 0 5 10 15 20 25 Element # +---------+---------+---------+---------+---------+ Na 35 òø Cl 38 òôòø V 29 ò÷ ùòòòòòòòòòø Mg 34 òø ó ó K 36 òôò÷ ó SO4 39 ò÷ ùòòòòòòòòòòòòòòòòòòòòòø Ni 20 òûòø ó ó Sr 27 ò÷ ùòø ó ó Ca 33 òòò÷ ùòòòòòòò÷ ó Co 6 òòòûò÷ ó Li 16 òòò÷ ùòòòòòòòòòòòòòø Ba 4 òø ó ó Ga 12 òôòòòòòòòòòòòòòòòòòòòø ó ó Eu 11 ò÷ ó ó ó Ge 14 òòòòòòòòòûòòòø ó ó ó U 28 òòòòòòòòò÷ ùòòòø ùòòòòòòòòòòòòò÷ ó Pb 22 òòòòòòòûòòòø ó ó ó ó pH 45 òòòòòòò÷ ùò÷ ó ó ó Ag 1 òòòòòòòòòûò÷ ó ó ó EC 43 òòòòòòòòò÷ ùòòò÷ ó Cs 8 òòòûòø ó ó Rb 24 òòò÷ ùòø ó ó Mo 18 òòòòò÷ ùòòòòòø ó ó As 3 òòòòòòò÷ ùòòò÷ ó Zr 32 òòòòòòòûòø ó ó NH4 41 òòòòòòò÷ ùòòò÷ ó Al 2 òòòòòòòòò÷ ó Mn 17 òòòûòø ó Sc 25 òòò÷ ùòòòø ó Fe 40 òòòòò÷ ó ó P 21 òòòûòø ùòòòòòòòòòòòòòòòòòø ó HCO3 37 òòò÷ ùòø ó ó ó Cr 7 òòòòò÷ ùò÷ ó ó NO3 42 òòòòòòò÷ ó ó Ce 5 òø ùòòòòòòòòòòòòòòòòòòòòò÷ Nd 19 òôòòòø ó La 15 ò÷ ùòòòø ó Temp 44 òòòòò÷ ó ó Er 10 òûòòòø ùòòòòòòòòòòòòòòòòò÷ Yb 31 ò÷ ó ó Gd 13 òòòø ùòòò÷ Y 30 òòòôòú Pr 23 òòò÷ ó Dy 9 òòòûò÷ Sm 26 òòò÷

Pear

sons

Cor

rela

tions

bet

wee

n el

emen

ts in

the

Seaw

ater

Intr

usio

n G

roun

dwat

ers

Ag

Al

As

Ba

Ce

Co

Cr

Cs

Dy

ErEu

Ga

Gd

Ge

LaLi

Mn

Mo

Nd

Ni

PPb

PrR

bSc

SmSr

UV

YYb

ZrC

aM

gN

aK

HC

O3

Cl

SO4

FeN

H4

NO

3EC

Tem

ppH

Ag

1.00

-0.0

60.

260.

110.

350.

080.

380.

160.

280.

380.

160.

080.

24-0

.08

0.37

0.23

0.29

0.07

0.30

0.15

0.13

-0.1

80.

48-0

.04

0.25

0.21

0.21

0.23

0.19

0.32

0.45

0.30

0.19

0.17

0.20

0.18

0.36

0.18

0.19

0.44

0.18

0.05

0.27

0.29

0.04

Al

1.00

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00.

30-0

.25

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60.

000.

09-0

.16

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20.

300.

35-0

.18

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7-0

.26

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3-0

.07

0.04

-0.1

70.

03-0

.11

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9-0

.29

0.14

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3-0

.07

-0.0

80.

070.

10-0

.13

-0.0

70.

31-0

.07

0.13

0.08

0.11

-0.1

30.

080.

11-0

.39

0.17

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1-0

.08

-0.4

00.

07A

s1.

000.

080.

430.

720.

290.

720.

330.

430.

120.

010.

47-0

.51

0.52

0.50

0.29

0.55

0.37

0.61

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4-0

.18

0.61

0.60

0.08

0.25

0.64

0.11

0.57

0.51

0.38

0.23

0.55

0.52

0.54

0.54

0.07

0.53

0.67

0.34

0.42

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10.

170.

32-0

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Ba

1.00

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50.

130.

180.

30-0

.32

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60.

971.

00-0

.24

-0.5

2-0

.11

0.13

0.48

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0-0

.13

0.44

0.47

0.25

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50.

330.

34-0

.14

0.38

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00.

39-0

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50.

090.

500.

470.

410.

360.

360.

430.

370.

420.

52-0

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0.53

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20.

39C

e1.

000.

630.

710.

340.

800.

82-0

.04

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90.

88-0

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0.97

0.72

0.66

0.07

0.98

0.54

0.36

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20.

920.

050.

620.

890.

63-0

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0.49

0.90

0.77

0.32

0.55

0.42

0.48

0.42

0.45

0.46

0.45

0.52

0.36

0.56

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20.

83-0

.04

Co

1.00

0.39

0.73

0.35

0.54

0.24

0.08

0.64

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80.

700.

910.

510.

540.

650.

89-0

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00.

650.

670.

210.

500.

900.

170.

880.

750.

490.

320.

760.

820.

860.

810.

100.

870.

860.

350.

48-0

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0.14

0.44

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6C

r1.

000.

070.

660.

710.

200.

140.

66-0

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0.65

0.50

0.69

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40.

740.

440.

65-0

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0.71

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60.

650.

670.

51-0

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0.47

0.69

0.68

0.36

0.46

0.43

0.47

0.43

0.80

0.44

0.43

0.47

0.61

0.62

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70.

460.

19C

s1.

000.

140.

320.

390.

250.

34-0

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0.45

0.60

0.27

0.57

0.34

0.73

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10.

040.

380.

83-0

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0.25

0.71

0.05

0.71

0.46

0.26

0.31

0.62

0.71

0.67

0.66

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90.

690.

750.

190.

35-0

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0.35

0.27

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y1.

000.

83-0

.27

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60.

860.

120.

720.

430.

340.

210.

790.

180.

24-0

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0.85

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20.

360.

870.

290.

010.

290.

790.

820.

440.

140.

250.

280.

310.

390.

230.

330.

190.

360.

51-0

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0.53

0.01

Er1.

00-0

.18

-0.3

10.

84-0

.04

0.75

0.62

0.38

0.33

0.84

0.40

0.13

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80.

870.

190.

270.

800.

480.

090.

530.

920.

970.

600.

290.

460.

510.

530.

350.

480.

520.

160.

370.

40-0

.20

0.48

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8Eu

1.00

0.97

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4-0

.61

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10.

230.

50-0

.07

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20.

520.

400.

21-0

.07

0.40

0.35

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40.

46-0

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0.47

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4-0

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0.13

0.56

0.54

0.49

0.42

0.30

0.51

0.44

0.44

0.47

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30.

61-0

.05

0.44

Ga

1.00

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8-0

.50

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60.

090.

44-0

.14

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70.

390.

450.

26-0

.21

0.30

0.32

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60.

32-0

.29

0.35

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1-0

.29

0.07

0.45

0.43

0.37

0.32

0.32

0.38

0.32

0.37

0.47

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00.

52-0

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0.39

Gd

1.00

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40.

820.

680.

530.

180.

890.

550.

21-0

.56

0.87

0.14

0.41

0.89

0.61

0.04

0.54

0.92

0.82

0.43

0.50

0.42

0.52

0.45

0.39

0.51

0.47

0.30

0.27

0.45

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40.

61-0

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Ge

1.00

-0.2

2-0

.43

-0.4

0-0

.15

-0.2

0-0

.73

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90.

09-0

.14

-0.5

4-0

.21

-0.0

6-0

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0.19

-0.5

8-0

.24

-0.0

1-0

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-0.7

0-0

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6-0

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0.01

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9-0

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6-0

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0.38

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7-0

.07

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9La

1.00

0.78

0.68

0.15

0.94

0.62

0.33

-0.5

00.

900.

120.

620.

790.

72-0

.04

0.53

0.86

0.70

0.32

0.63

0.48

0.52

0.46

0.41

0.51

0.51

0.60

0.42

0.48

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40.

88-0

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Li1.

000.

620.

440.

760.

850.

13-0

.48

0.72

0.52

0.34

0.59

0.90

0.17

0.88

0.81

0.59

0.38

0.74

0.84

0.89

0.82

0.28

0.89

0.82

0.46

0.51

0.14

0.10

0.56

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n1.

00-0

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0.66

0.71

0.78

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10.

55-0

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0.90

0.48

0.76

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50.

500.

520.

360.

230.

840.

460.

520.

360.

760.

520.

390.

820.

530.

470.

040.

660.

30M

o1.

000.

080.

27-0

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10.

310.

80-0

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0.06

0.30

0.55

0.58

0.35

0.34

0.38

0.00

0.62

0.56

0.71

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50.

540.

70-0

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0.41

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60.

20-0

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6N

d1.

000.

580.

36-0

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0.89

0.08

0.60

0.89

0.66

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50.

550.

920.

800.

410.

560.

480.

540.

480.

440.

520.

490.

470.

370.

55-0

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0.77

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4N

i1.

000.

23-0

.28

0.51

0.59

0.40

0.40

0.98

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70.

880.

630.

350.

280.

950.

820.

870.

740.

290.

890.

800.

510.

49-0

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0.17

0.41

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1.00

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00.

27-0

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0.88

0.28

0.29

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10.

050.

140.

100.

060.

440.

070.

08-0

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0.88

0.07

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10.

700.

470.

66-0

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0.44

0.48

Pb1.

00-0

.58

0.08

-0.3

0-0

.50

-0.3

1-0

.25

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1-0

.53

-0.5

4-0

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-0.2

2-0

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-0.3

1-0

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9-0

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7-0

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5-0

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0.30

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70.

33Pr

1.00

0.18

0.47

0.81

0.62

0.12

0.57

0.91

0.85

0.51

0.47

0.51

0.56

0.55

0.46

0.53

0.58

0.49

0.47

0.48

-0.0

70.

70-0

.13

Rb

1.00

-0.3

60.

060.

550.

290.

770.

310.

170.

300.

370.

790.

740.

80-0

.34

0.75

0.82

-0.1

00.

39-0

.58

0.40

-0.1

5-0

.25

Sc1.

000.

460.

47-0

.57

0.14

0.35

0.24

0.10

0.62

0.12

0.16

0.02

0.78

0.16

0.07

0.83

0.36

0.68

-0.0

30.

730.

36Sm

1.00

0.45

0.05

0.45

0.85

0.77

0.39

0.36

0.39

0.45

0.42

0.38

0.42

0.42

0.24

0.32

0.51

-0.2

30.

610.

00Sr

1.00

-0.1

00.

870.

700.

430.

310.

940.

820.

870.

750.

360.

880.

810.

600.

550.

090.

150.

52-0

.08

U1.

000.

210.

120.

130.

14-0

.28

0.19

0.22

0.30

-0.3

40.

200.

22-0

.36

-0.0

6-0

.27

0.17

-0.2

1-0

.52

V1.

000.

700.

510.

480.

700.

971.

000.

970.

210.

990.

960.

280.

62-0

.10

0.21

0.21

-0.1

9Y

1.00

0.88

0.50

0.53

0.62

0.69

0.65

0.34

0.67

0.66

0.30

0.38

0.35

-0.1

70.

58-0

.13

Yb1.

000.

630.

240.

440.

490.

520.

330.

460.

490.

140.

350.

37-0

.18

0.41

-0.0

3Zr

1.00

0.17

0.47

0.46

0.53

0.19

0.45

0.48

0.04

0.41

0.04

0.00

0.04

0.01

Ca

1.00

0.64

0.70

0.53

0.43

0.73

0.61

0.71

0.44

0.13

0.16

0.54

0.08

Mg

1.00

0.98

0.98

0.18

0.97

0.97

0.28

0.72

-0.1

50.

310.

18-0

.19

Na

1.00

0.96

0.23

1.00

0.95

0.30

0.63

-0.0

80.

230.

21-0

.18

K1.

000.

150.

950.

970.

180.

70-0

.14

0.26

0.12

-0.2

5H

CO

31.

000.

210.

120.

660.

510.

68-0

.09

0.44

0.34

Cl

1.00

0.94

0.31

0.60

-0.1

10.

240.

21-0

.17

SO4

1.00

0.24

0.69

-0.1

80.

250.

17-0

.23

Fe1.

000.

370.

430.

310.

700.

25N

H4

1.00

0.15

0.19

0.21

0.01

NO

31.

00-0

.35

0.56

0.09

EC1.

000.

020.

23Te

mp

1.00

-0.0

7pH

1.00

Alth

ough

cor

rela

tions

bet

wee

n ar

seni

c an

d ot

her e

lem

ents

are

hig

hlig

hted

her

e, o

ther

sig

nific

ant c

orre

latio

ns e

xist

bet

wee

en o

ther

ele

men

ts in

thes

e w

ater

s. T

hese

oth

er c

orre

latio

ns a

re c

urre

ntly

the

topi

c of

furt

her s

tudy

by

the

auth

or h

erei

n.

App

endi

x B

7 - S

tatis

tical

Res

ults

SHA

LLO

W G

RO

UN

DW

ATE

RS

SELE

CTE

D G

EOC

HEM

ICA

L M

OD

ELLI

NG

OU

TPU

T

Arsenic

Species

Molalities

SAM

PLE

solu

tion

m_H

2AsO

3-m

_H2A

sO4-

m_H

3AsO

3m

_H3A

sO4

m_H

AsO

3-2

m_H

AsO

4-2

As(

III)

As(

V)To

t As

As(

III)/Tot

As

ML1/5

15.70E-

122.

64E-

123.

52E-

081.

13E-

151.71E-

181.

17E-

133.

52E-

082.75E-

123.

52E-

0810

0ML1/6

23.42E-

121.

07E-

143.

51E-

087.

81E-

186.

50E-

192.

94E-

163.

51E-

081.

10E-

143.

51E-

0810

0ML1/7

38.39E-

125.47E-

144.73E-

082.

18E-

172.79E-

182.

67E-

154.73E-

085.74E-

144.73E-

0810

0ML2/5

42.39E-

124.

11E-

182.

57E-

083.71E-21

4.47E-

199.72E-20

2.58E-

084.21E-

182.

58E-

0810

0ML2/6

52.28E-

111.

07E-

152.71E-

071.

04E-

183.

81E-

182.30E-

172.71E-

071.

09E-

152.71E-

0710

0ML2/7

63.

10E-

111.

66E-

154.

11E-

071.

82E-

184.

60E-

183.

14E-

174.

11E-

071.

69E-

154.

11E-

0710

0ML2/8

72.43E-

117.

54E-

162.45E-

076.

12E-

194.

65E-

181.

90E-

172.45E-

077.74E-

162.45E-

0710

0ML3/3

81.

1 3E-

111.

65E-

171.

12E-

071.

14E-20

2.15E-

184.

65E-

191.

12E-

071.70E-

171.

12E-

0710

0ML4/3

94.36E-

114.41E-

162.

04E-

071.45E-

191.

85E-

172.74E-

172.

04E-

074.

68E-

162.

04E-

0710

0ML4/4

102.

62E-

124.30E-

172.

61E-

083.

02E-20

5.00E-

191.20E-

182.

61E-

084.42E-

172.

61E-

0810

0ML4/5

117.

07E-

121.21E-

154.

90E-

085.

94E-

191.

93E-

184.

84E-

174.

90E-

081.26E-

154.

90E-

0810

0ML4/6

121.30E-

125.36E-

173.

58E-

081.

04E-

199.

00E-20

5.42E-

193.

58E-

085.42E-

173.

58E-

0810

0ML5/4

136.78E-

141.46E-

142.

54E-

094.

00E-

173.34E-21

1.02E-

162.

54E-

091.47E-

142.

54E-

0910

0ML5/5

148.31E-

131.

15E-

149.

93E-

091.

02E-

171.31E-

192.

56E-

169.

93E-

091.

18E-

149.

93E-

0910

0ML5/6

152.41E-

123.

58E-

141.22E-

081.33E-

179.

10E-

191.

92E-

151.22E-

083.77E-

141.22E-

0810

0ML6/7

165.24E-

131.

03E-

142.30E-

083.20E-

172.23E-20

6.42E-

172.30E-

081.

04E-

142.30E-

0810

0ML6/8

171.47E-

121.78E-

144.

17E-

083.

60E-

179.

99E-20

1.77E-

164.

17E-

081.

80E-

144.

17E-

0810

0ML6/9

186.

82E-

131.37E-

131.

05E-

081.

50E-

169.30E-20

2.73E-

151.

05E-

081.40E-

131.

05E-

0810

0ML6/1

019

3.28E-

125.

88E-

132.42E-

083.

10E-

169.

14E-

192.38E-

142.42E-

086.

12E-

132.42E-

0810

0

Satu

ration Indices

SAM

PLE

solu

tion

si_A

l(OH

)3(a

m)

si_A

l2O

3si

_Al4

(OH

)10S

O4

si_A

lOH

SO4

si_Boe

hmite

si_D

iaspor

esi

_Gib

bsite

si_Goe

thite

si_F

errih

ydrit

esi

_Hem

atite

si_Orp

imen

tsi

_Pyr

itesi

_Rea

lgar

ML1/5

1-1

.101

9-0

.37

1.4941

-2.7

057

1.09

932.

8498

1.45

56-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99ML1/6

2-1

.2547

-0.6

643

1.8927

-1.7477

0.94

862.

6947

1.29

81-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99ML1/7

3-0

.553

60.7334

4.3124

-1.4717

1.64

893.39

672.

0011

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

ML2/5

4-1

.128

8-0

.337

92.

1933

-1.145

91.

0881

2.8044

1.3922

2.5321

-0.177

87.45

82-9

99.9

99-9

99.9

99-9

99.9

99ML2/6

5-1

.5874

-1.2

661

0.5625

-1.4

996

0.6275

2.34

810.

9382

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

ML2/7

6-1

.701

5-1

.49

0.17

98-1

.5003

0.5141

2.233

0.8222

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

ML2/8

7-1

.103

9-0

.3077

2.3943

-1.1

977

1.10

942.

8335

1.42

55-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99ML3/3

8-0

.699

10.43

564.

037

-1.3712

1.5021

3.2526

1.85

84-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99ML4/3

9-0

.4414

0.95

575.

1817

-0.9

593

1.76

073.

5094

2.1143

3.2931

0.5548

8.9628

-999

.999

-999

.999

-999

.999

ML4/4

10-0

.815

60.20

933.

5444

-1.4

535

1.38

683.

1347

1.73

91-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99ML4/5

11-0

.611

0.61

863.

6085

-2.0

033

1.59

143.33

931.

9437

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

ML4/6

12-1

.624

1-1

.4077

0.81

58-1

.756

60.

5783

2.32

620.

9306

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

ML5/4

13-2

.2763

-2.6

943

-1.8

915

-2.346

1-0

.070

61.

6702

0.27

09-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99ML5/5

14-1

.520

5-1

.180

50.

6815

-2.0203

0.68

562.42

541.

0257

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

ML5/6

15-2

.2043

-2.5

502

-2.4324

-3.1

031

0.00

141.7422

0.3429

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

ML6/7

16-2

.8284

-3.8

163

-3.2

63-2

.222

5-0

.626

1.12

18-0

.273

8-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99ML6/8

17-1

.771

9-1

.701

11.

141

-0.9

677

0.43

092.

1779

0.78

18-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99ML6/9

18-1

.3077

-0.772

63.

0385

-0.4

629

0.89

512.

6421

1.24

6-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99-9

99.9

99ML6/1

019

-1.0

988

-0.3

526

3.1739

-0.933

91.

1044

2.85

051.4539

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

-999

.999

App

endi

x B8

- PH

REE

QC

Out

put

BA

RR

IER

SA

ND

GR

OU

ND

WA

TER

S SE

LEC

TED

GEO

CH

EMIC

AL

OU

TPU

T

Arsenic

Species

Molalities

SAM

PLE

solu

tion

m_A

sO3-

3m

_AsO

4-3

m_H

2AsO

3-m

_H2A

sO4-

m_H

3AsO

3m

_H3A

sO4

m_H

AsO

3-2

m_H

AsO

4-2

As(

III)

As(

V)To

t As

As(

III)/Tot

As

ML1/6

13.31E-20

3.12E-

161.

03E-

098.20E-

138.44E-

085.

04E-

182.

58E-

142.

84E-

128.

54E-

083.

66E-

128.

54E-

0810

0ML1/7

27.

50E-20

1.19E-

151.

62E-

092.

19E-

121.

06E-

071.

08E-

174.

93E-

149.

19E-

121.

08E-

071.

14E-

111.

08E-

0710

0ML 1/8

34.26E-

198.

19E-

152.

12E-

093.

52E-

126.

86E-

088.

69E-

181.34E-

133.

02E-

117.

07E-

083.37E-

117.

08E-

0810

0ML2/6

41.26E-

191.43E-

142.

59E-

092.

53E-

111.

68E-

071.25E-

168.

04E-

141.

08E-

101.71E-

071.33E-

101.71E-

0710

0ML2/7

53.

01E-20

1.91E-

151.

51E-

098.

13E-

121.

56E-

076.32E-

172.

99E-

142.22E-

111.

57E-

073.

03E-

111.

58E-

0710

0ML2/8

61.

92E-20

3.53E-

151.22E-

092.

07E-

111.47E-

072.

04E-

162.

14E-

144.

68E-

111.48E-

076.75E-

111.48E-

0710

0ML2/9

74.

03E-20

9.72E-

151.73E-

093.79E-

111.

68E-

072.

97E-

163.

69E-

141.

06E-

101.

69E-

071.44E-

101.70E-

0710

0ML2/1

08

5.19E-20

6.85E-

152.

59E-

092.

84E-

112.

68E-

072.20E-

165.

10E-

147.

85E-

112.71E-

071.

07E-

102.71E-

0710

0ML3/5

.59

7.96E-21

2.60E-

174.

80E-

101.

16E-

135.43E-

088.

81E-

198.

53E-

153.

17E-

135.47E-

084.33E-

135.47E-

0810

0ML3/6

.510

2.93E-21

3.75E-

182.76E-

102.73E-

143.

84E-

082.

63E-

193.

96E-

155.

81E-

143.

87E-

088.

54E-

143.

87E-

0810

0ML3/7

.511

3.86E-22

2.97E-

198.

86E-

115.

04E-

151.

99E-

087.

55E-20

8.01E-

167.

00E-

152.

00E-

081.20E-

142.

00E-

0810

0ML3/8

.512

2.03E-21

9.55E-

181.76E-

106.

60E-

142.39E-

086.38E-

192.

62E-

151.43E-

132.40E-

082.

09E-

132.40E-

0810

0ML3/9

.513

5.93E-21

5.52E-

183.

64E-

102.

55E-

144.

10E-

081.

94E-

196.44E-

156.

84E-

144.

14E-

089.39E-

144.

14E-

0810

0ML3/1

0.5

141.39E-20

2.33E-

175.21E-

106.

64E-

144.

62E-

084.

02E-

191.

18E-

142.26E-

134.

67E-

082.

93E-

134.

67E-

0810

0ML4/ 5

155.

99E-20

8.73E-

111.21E-

102.

08E-

083.48E-

096.

15E-

141.

10E-

142.

03E-

073.

60E-

092.24E-

072.27E-

072

ML4/6

161.48E-21

3.22E-

172.

93E-

105.

55E-

137.

05E-

081.

03E-

172.76E-

157.

09E-

137.

08E-

081.26E-

127.

08E-

0810

0ML4/7

1 71.26E-21

5.94E-

182.

91E-

101.23E-

137.

05E-

082.35E-

182.

61E-

151.46E-

137.

08E-

082.

69E-

137.

08E-

0810

0ML4/9

181.38E-21

5.45E-

181.

81E-

106.

80E-

143.32E-

081.

04E-

182.

16E-

151.

03E-

133.34E-

081.71E-

133.34E-

0810

0ML4/1

019

2.79E-21

2.70E-

172.

15E-

102.

05E-

133.

05E-

082.

51E-

183.36E-

153.

94E-

133.

07E-

085.

99E-

133.

07E-

0810

0ML4/1

120

1.64E-21

1.59E-

172.

01E-

101.

91E-

133.

58E-

082.

95E-

182.49E-

152.

92E-

133.

60E-

084.

83E-

133.

61E-

0810

0ML4/12

211.27E-21

3.08E-

181.

56E-

103.73E-

142.79E-

085.76E-

191.

9 3E-

155.

69E-

142.

80E-

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221.

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232.

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91A

ppen

dix

B8

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Out

put

FLU

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L SA

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LEC

TED

GEO

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Arsenic

Species

Molalities

SAM

PLE

Solu

tion

m_A

sO3-

3m

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m_H

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3-m

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sO4-

m_H

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As(

III)

As(

V)To

t As

As(

III)/Tot

As

ML1/9

13.

82E-20

3.35E-

165.41E-

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142.

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122.

97E-

0810

0ML1/1

02

1.63E-

197.

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58E-

181.

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076.28E-

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00E-

0710

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13

4.30E-20

7.85E-

157.

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101.

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114.

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084.

66E-

172.45E-

145.22E-

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086.30E-

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0810

0ML1/12

43.37E-

194.33E-

132.

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093.42E-

101.22E-

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151.41E-

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ML1/13

56.37E-2 0

3.84E-

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66E-

104.

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123.74E-

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173.

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142.24E-

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082.

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113.

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0810

0ML1/14

61.

56E-

194.

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093.

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126.76E-

081.30E-

176.

91E-

142.35E-

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082.73E-

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0810

0ML1/1

57

2.60E-20

6.58E-

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89E-

101.

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088.

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181.

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145.

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087.20E-

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0810

0ML1/1

68

6.36E-20

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183.73E-

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126.

56E-

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116.

56E-

0810

0ML2/1

19

7.79E-20

2.16E-

142.

05E-

094.

91E-

111.

52E-

072.78E-

165.

62E-

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101.

54E-

072.33E-

101.

5 4E-

0710

0ML2/12

101.

55E-

192.

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092.41E-

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07E-

079.

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178.

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141.32E-

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071.

56E-

101.

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0710

0ML2/13

117.

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191.43E-

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071.77E-

162.

53E-

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84E-

101.

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56E-

101.

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0710

0ML2/14

126.78E-

198.

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162.46E-

133.72E-

101.

64E-

074.

18E-

101.

64E-

0710

0ML2/1

513

3.16E-

191.47E-

142.

95E-

091.

06E-

111.26E-

073.

13E-

171.36E-

137.25E-

111.29E-

078.31E-

111.29E-

0710

0ML2/1

614

1.68E-

198.41E-

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091.

0 4E-

111.37E-

074.26E-

179.

08E-

145.31E-

111.39E-

076.35E-

111.39E-

0710

0ML2/17

151.

03E-

196.

98E-

141.

68E-

091.

04E-

109.

68E-

084.

85E-

165.

91E-

144.75E-

109.

85E-

085.79E-

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91E-

0899

ML2/1

816

7.04E-20

4.70E-

151.

67E-

099.20E-

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074.73E-

174.

80E-

143.74E-

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074.

66E-

111.

18E-

0710

0ML3/12.

517

4.11E-21

1.77E-

172.

83E-

109.30E-

143.43E-

087.73E-

194.71E-

152.32E-

133.46E-

083.25E-

133.46E-

0810

0ML3/13.

518

6.40E-21

8.21E-

183.75E-

103.

67E-

144.25E-

082.

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196.74E-

159.

89E-

144.29E-

081.36E-

134.29E-

0810

0ML3/14.

519

5.53E-21

3.41E-

173.77E-

101.78E-

134.

66E-

081.

51E-

186.26E-

154.42E-

134.70E-

086.20E-

134.70E-

0810

0ML3/1

5.5

206.

05E-21

2.54E-

174.

55E-

101.48E-

135.

87E-

081.32E-

187.21E-

153.48E-

135.

91E-

084.

95E-

135.

91E-

0810

0ML3/ 1

6.5

218.

06E-21

5.52E-

176.27E-

103.33E-

138.22E-

083.

03E-

189.79E-

157.

69E-

138.28E-

081.

10E-

128.28E-

0810

0ML3/17.

522

1.28E-20

2.36E-

178.33E-

101.

19E-

131.

01E-

079.

99E-

191.42E-

143.

00E-

131.

01E-

074.

19E-

131.

01E-

0710

0ML3/1

8.5

231.

00E-20

6.6 4E-

178.22E-

104.

19E-

131.

11E-

073.

89E-

181.25E-

149.48E-

131.

12E-

071.37E-

121.

12E-

0710

0ML3/1

9.5

241.

59E-20

1.90E-

171.

12E-

091.

03E-

131.42E-

078.

97E-

191.

83E-

142.

51E-

131.43E-

073.

55E-

131.43E-

0710

0ML1/1

925

2.60E-

195.

80E-

156.

06E-

091.22E-

114.31E-

076.

92E-

171.76E-

134.

66E-

114.37E-

075.

88E-

114.37E-

0710

0ML1/2

026

2.07E-

196.26E-

156.

08E-

091.

64E-

114.

85E-

071.

03E-

161.

57E-

135.

63E-

114.

91E-

077.27E-

114.

91E-

0710

0ML1/1

127

8.63E-21

2.26E-

183.70E-

107.

63E-

153.38E-

084.

93E-20

7.95E-

152.40E-

143.42E-

083.

16E-

143.42E-

0810

0ML 1/12

281.37E-20

1.62E-

164.

87E-

104.

59E-

134.

11E-

082.76E-

181.

15E-

141.

57E-

124.

16E-

082.

03E-

124.

16E-

0810

0ML 1/13

291.33E-20

1.18E-

164.29E-

103.

06E-

133.

50E-

081.

80E-

181.

06E-

141.

09E-

123.

54E-

081.39E-

123.

54E-

0810

0ML1/14

307.35E-21

2.44E-

163.

86E-

101.

03E-

124.

06E-

087.71E-

187.42E-

152.

85E-

124.

10E-

083.

87E-

124.

10E-

0810

0ML1/1

531

1.39E-20

1.99E-

174.72E-

105.46E-

143.

99E-

083.34E-

191.

13E-

141.

87E-

134.

04E-

082.42E-

134.

04E-

0810

0ML1/1

632

7.53E-21

7.45E-

184.39E-

103.

52E-

144.

90E-

082.

83E-

198.

00E-

159.

18E-

144.

94E-

081.27E-

134.

94E-

0810

0ML1/17

335.43E-21

3.39E-

184.

16E-

102.

10E-

145.32E-

081.

9 4E-

196.

61E-

154.78E-

145.36E-

086.

88E-

145.36E-

0810

0ML1/1

834

1.07E-20

1.93E-

177.72E-

101.

13E-

139.

63E-

081.

01E-

181.26E-

142.

64E-

139.71E-

083.76E-

139.71E-

0810

0ML1/1

935

2.13E-20

1.48E-

161.47E-

098.33E-

131.79E-

077.36E-

182.46E-

141.

99E-

121.

80E-

072.

82E-

121.

80E-

0710

0ML1/2

036

6.28E-20

1.03E-

153.

55E-

094.

81E-

123.

89E-

073.

90E-

176.

59E-

141.25E-

113.

92E-

071.74E-

113.

92E-

0710

0ML2/1

137

4.25E-21

1.31E-

162.44E-

107.

03E-

132.

80E-

086.

61E-

184.

50E-

151.

66E-

122.

82E-

082.37E-

122.

82E-

0810

0ML2/12

385.

08E-21

6.00E-

172.21E-

102.30E-

132.20E-

081.78E-

184.

66E-

156.

50E-

132.22E-

088.

80E-

132.22E-

0810

0ML2/13

394.44E-21

7.33E-

172.

51E-

103.

62E-

132.78E-

083.

09E-

184.

67E-

159.

10E-

132.

80E-

081.27E-

122.

80E-

0810

0ML2/14

406.

00E-21

5.05E-

184.

84E-

103.

56E-

146.

63E-

083.79E-

197.45E-

157.40E-

146.

68E-

081.

10E-

136.

68E-

0810

0ML2/1

541

1.04E-20

4.29E-

196.

53E-

102.27E-

157.

83E-

082.

04E-20

1.14E-

145.48E-

157.

90E-

087.75E-

157.

90E-

0810

0ML2/ 1

642

5.18E-21

6.10E-

186.41E-

106.

54E-

141.

05E-

078.24E-

198.

05E-

151.

12E-

131.

06E-

071.77E-

131.

06E-

0710

0ML2/17

434.44E-23

1.22E-20

5.38E-

111.28E-

152.77E-

085.

11E-20

2.16E-

166.

96E-

162.78E-

081.

98E-

152.78E-

0810

0ML2/ 1

844

9.24E-21

1.45E-

171.

19E-

091.

61E-

132.

01E-

072.

06E-

181.46E-

142.70E-

132.

02E-

074.31E-

132.

02E-

0710

0ML2/1

945

4.95E-21

2.46E-

189.

93E-

104.23E-

142.

10E-

076.

82E-

199.73E-

155.

68E-

142.

11E-

079.

91E-

142.

11E-

0710

0ML2/23

465.75E-20

5.42E-

151.

98E-

091.

58E-

111.70E-

071.

02E-

164.71E-

145.20E-

111.72E-

076.78E-

111.72E-

0710

0ML2/24

474.

84E-20

6.68E-

151.

85E-

092.

13E-

111.

67E-

071.43E-

164.

17E-

146.71E-

111.

69E-

078.

84E-

111.

69E-

0710

0ML2/2

548

3.67E-20

2.03E-

151.

68E-

097.

56E-

121.

66E-

075.43E-

173.45E-

142.22E-

111.

68E-

072.

97E-

111.

68E-

0710

0ML2/2

649

4.79E-20

7.03E-

151.

67E-

092.

18E-

111.45E-

071.48E-

163.

96E-

146.

89E-

111.46E-

079.

07E-

111.46E-

0710

0ML2/27

501.79E-

194.37E-

135.30E-

091.

17E-

094.23E-

077.46E-

151.37E-

133.

97E-

094.29E-

075.

14E-

094.34E-

0799

ML2/2

851

8.35E-

192.

11E-

121.27E-

082.71E-

097.

18E-

071.

16E-

144.

55E-

131.35E-

087.31E-

071.

62E-

087.47E-

0798

ML2/2

952

2.88E-

192.45E-

125.

83E-

094.

06E-

093.

80E-

071.

93E-

141.

81E-

131.79E-

083.

86E-

072.

19E-

084.

08E-

0795

ML2/2

553

1.98E-20

1.2 0E-

161.

95E-

091.

00E-

122.78E-

071.

08E-

172.76E-

141.

96E-

122.

80E-

072.

96E-

122.

80E-

0710

0ML2/2

654

1.01E-20

6.71E-

171.33E-

097.

64E-

132.21E-

079.70E-

181.

6 3E-

141.27E-

122.22E-

072.

04E-

122.22E-

0710

0ML2/27

551.

67E-20

8.49E-

171.72E-

097.

61E-

132.

50E-

078.

63E-

182.38E-

141.43E-

122.

52E-

072.

19E-

122.

52E-

0710

0ML6/13

562.

50E-21

3.26E-

122.30E-

112.43E-

091.

11E-

098.47E-

151.

02E-

151.

55E-

081.

13E-

091.79E-

081.

90E-

086

ML 6/14

571.30E-20

4.52E-

129.

60E-

112.

99E-

094.

12E-

091.

02E-

144.

84E-

151.

99E-

084.22E-

092.29E-

082.72E-

0816

ML6/1

558

8.04E-21

2.65E-

121.

00E-

102.

88E-

095.

57E-

091.23E-

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89E-

151.

51E-

085.

67E-

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80E-

082.37E-

0824

ML6/1

659

1.39E-20

1.33E-

133.

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102.

60E-

102.28E-

081.

51E-

158.

97E-

151.

01E-

092.31E-

081.27E-

092.44E-

0895

ML6/17

601.75E-20

6.09E-

132.

11E-

106.

59E-

101.

17E-

082.

89E-

158.32E-

153.43E-

091.

19E-

084.

09E-

091.

60E-

0874

App

endi

x B8

- PH

REE

QC

Out

put

FLU

VIA

L SA

ND

GR

OU

ND

WA

TER

S SE

LEC

TED

GEO

CH

EMIC

AL

OU

TPU

T

Satu

ration Indices

SAM

PLE

Solu

tion

si_F

e3(O

H)8

si_A

gmet

alsi

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hmite

si_F

errih

ydrit

esi

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atite

si_Cr2

O3

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283.78E-21

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291.22E-20

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332.30E-21

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935

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538

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639

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144

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950

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155

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561.

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576.

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582.

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559

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660

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611.

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862

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963

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165

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675.

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681.

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670

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712.45E-20

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872

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ML6/1

852

-999

.999

-4.044

50.

1331

-999

.999

-999

.999

1.9021

0.5673

1.8642

1.7779

3.9828

0.45

98-9

99.9

99-9

99.9

994.2072

-999

.999

-1.7482

11.4

553

14.237

5-3

.4533

0.43

94ML6/1

953

-0.6

029

-3.527

80.2077

2.1624

12.1

652

1.3633

0.31

021.

9458

0.183

3.08

110.

5451

4.89

1417

.929

54.

9105

4.14

84-2

.486

12.3752

12.078

9-4

.477

-1.3332

ML6/2

054

-0.231

6-3

.374

61.

0598

2.3441

12.529

51.

5238

0.392

2.79

880.3372

4.77

981.39

865.

0738

17.6

163.

6898

4.3345

-2.2

8512

.588

612

.2187

-4.1

509

-1.0253

ML6/2

155

-999

.999

-3.1347

0.56

61-9

99.9

99-9

99.9

991.36

130.3138

2.30

680.7228

4.34

170.

9075

-999

.999

-999

.999

3.12

14-9

99.9

99-1

.6404

11.7267

12.6467

-2.3

075

-0.7

557

ML6/22

56-9

99.9

99-3

.5262

0.45

13-9

99.9

99-9

99.9

991.444

0.35

822.

1939

0.25

133.

5595

0.79

55-9

99.9

99-9

99.9

992.3375

-999

.999

-2.1724

12.4

073

12.8

093

-3.9

164

-0.373

8ML6/23

57-9

99.9

99-1

.771

91.4525

-999

.999

-999

.999

2.13

180.7036

3.19

592.

0103

6.6336

1.7979

-999

.999

-999

.999

4.1227

-999

.999

-1.449

910

.6904

13.7

993

-2.3277

0.46

64ML6/24

58-9

99.9

99-3

.814

8-0

.4-9

99.9

99-9

99.9

992.

0846

0.68

151.3443

1.56

922.

5354

-0.0

532

-999

.999

-999

.999

3.7543

-999

.999

-1.7

108

11.9

978

14.1

115

-2.8

549

0.6643

ML6/2

559

-999

.999

-3.8273

0.90

67-9

99.9

99-9

99.9

991.73

930.

5089

2.65

10.

9938

4.91

861.2535

-999

.999

-999

.999

3.0175

-999

.999

-1.770

912

.5897

13.9

116

-2.5

661

0.44

94ML6/2

660

-999

.999

-1.636

1.33

16-9

99.9

99-9

99.9

991.

0112

0.1448

3.07

59-0

.2343

5.26

861.

6785

-999

.999

-999

.999

1.54

08-9

99.9

99-2

.057

86.

159

9.30

64-2

.983

8-1

.6333

ML6/27

61-9

99.9

99-3

.538

1.48

18-9

99.9

99-9

99.9

991.30

070.2926

3.2279

0.31

945.

8346

1.83

14-9

99.9

99-9

99.9

991.75

09-9

99.9

99-1

.840

112

.5997

13.142

-2.2

862

-0.3

08ML6/2

862

-999

.999

-3.573

11.

0071

-999

.999

-999

.999

0.95

010.

139

2.76

56-0

.6417

4.2841

1.37

56-9

99.9

99-9

99.9

990.35

63-9

99.9

99-2

.2224

12.524

812

.948

1-3

.217

5-0

.2332

ML6/2

963

-999

.999

-3.6397

1.18

69-9

99.9

99-9

99.9

991.4274

0.36

812.

9401

0.3026

5.10

671.

5471

-999

.999

-999

.999

1.5625

-999

.999

-1.742

110

.877

13.4377

-1.7783

0.11

97ML6/3

064

-999

.999

-1.425

80.4242

-999

.999

-999

.999

1.50

950.40

072.

1729

0.93

14.

1247

0.7771

-999

.999

-999

.999

2.88

08-9

99.9

99-1

.082

812

.9337

10.826

8-0

.5727

-3.3

93ML1/2

165

3.0149

2.2735

-1.223

93.46

6314

.76

1.6836

0.452

0.5037

1.6837

1.39

06-0

.902

56.

186

9.51

94.

1146

5.39

94-0

.9922

5.1546

9.0347

-0.649

8-2

.438

5ML1/22

662.

8479

8.3447

-1.279

93.47

1414

.759

82.

1003

0.64

530.43

892.

0641

1.23

54-0

.971

96.

1836

-13.

5052

4.39

965.36

08-1

.0613

-11.

9875

-2.5

575

-0.8

856

-8.1

871

ML1/23

673.

0346

2.27

61-0

.583

13.4624

14.741

2.232

0.70

941.

1349

2.4146

2.8474

-0.276

56.

1739

9.27

124.

6839

5.3476

-0.7

674.

8652

8.9877

-0.037

1-2

.498

9ML1/24

682.74

821.79

16-1

.215

13.29

1814

.413

51.7475

0.48

680.

5143

1.80

181.4637

-0.8

912

6.0132

11.8

144.

9989

5.2338

-0.8262

6.2936

9.382

0.32

85-2

.8043

ML1/2

569

1.6839

0.34

02-1

.258

82.7938

13.4

197

1.75

140.49

180.4722

1.70

91.28

08-0

.9323

5.51

6812

.117

95.3028

4.7446

-0.9

875

6.5775

11.6

521

-0.040

6-0

.7982

ML1/2

670

3.02

512.

5428

-1.232

83.4272

14.6

844

1.7673

0.49

670.49

661.

8276

1.4343

-0.9

089

6.14

869.

9509

5.1378

5.36

92-0

.863

83.7639

8.43

80.43

82-3

.2193

ML1/27

712.

8845

3.05

50.

1524

3.5329

14.8747

1.9227

0.5443

1.8643

1.9121

4.12

050.44

986.23

917.76

194.

5185

5.3877

-0.936

81.

5602

7.56

10.

0125

-3.328

ML1/2

872

1.83

13-0

.199

9-1

.113

63.20

6414

.1894

1.5976

0.33

580.

5718

1.61

541.

5967

-0.8

567

5.88

9412

.058

63.

1987

4.92

81-0

.874

14.

9013

10.9

155

0.0576

-1.0

57

App

endi

x B8

- PH

REE

QC

Out

put

SEA

WA

TER

INTR

USI

ON

GR

OU

ND

WA

TER

S SE

LEC

TED

GEO

CH

EMIC

AL

OU

TPU

T

Arsenic

Species

Molalities

Sam

ple

Solu

tion

m_A

sO3-

3m

_AsO

4-3

m_H

2AsO

3-m

_H2A

sO4-

m_H

3AsO

3m

_H3A

sO4

m_H

AsO

3-2

m_H

AsO

4-2

As(

III)

As(

V)To

t As

As(

III)/Tot

As

ML1/2

61

1.59E-

192.

65E-20

1.93E-

093.

02E-

173.73E-

074.

63E-22

5.01E-

141.

00E-

163.75E-

071.30E-

163.75E-

0710

0ML1/27

21.

63E-

193.32E-

192.

11E-

093.

99E-

164.

60E-

076.79E-21

5.16E-

141.25E-

154.

62E-

071.

65E-

154.

62E-

0710

0ML1/2

83

8.87E-20

1.49E-

191.79E-

092.70E-

165.32E-

076.

08E-21

3.38E-

146.74E-

165.33E-

079.44E-

165.33E-

0710

0ML1/2

94

1.59E-

192.

65E-20

2.28E-

093.

14E-

177.71E-

077.22E-22

4.60E-

148.

91E-

177.74E-

071.20E-

167.74E-

0710

0ML1/3

05

2.08E-

195.

00E-20

2.45E-

094.

67E-

177.

97E-

079.

91E-22

5.30E-

141.47E-

167.

99E-

071.

94E-

167.

99E-

0710

0ML3/17

64.39E-

191.76E-

174.74E-

091.45E-

144.27E-

078.78E-20

1.61E-

137.41E-

144.32E-

078.

86E-

144.32E-

0710

0ML3/1

87

5.88E-

186.

11E-20

6.89E-

095.

50E-

185.38E-

072.78E-23

5.17E-

136.

16E-

175.45E-

076.71E-

175.45E-

0710

0ML3/1

98

1.41E-

176.

11E-

198.

98E-

092.

98E-

175.

85E-

071.25E-22

8.77E-

134.35E-

165.

94E-

074.

65E-

165.

94E-

0710

0ML3/2

09

6.71E-

186.47E-

196.

92E-

095.

16E-

174.

86E-

072.36E-22

5.62E-

136.22E-

164.

93E-

076.74E-

164.

93E-

0710

0ML3/2

110

1.98E-

189.

60E-

194.

16E-

091.

56E-

162.74E-

076.

85E-22

2.72E-

131.

52E-

152.78E-

071.

67E-

152.78E-

0710

0ML3/22

113.

52E-

184.70E-

185.

12E-

095.28E-

163.

00E-

072.

06E-21

3.94E-

136.

03E-

153.

05E-

076.

57E-

153.

05E-

0710

0ML3/23

127.46E-

187.34E-

187.32E-

095.

59E-

163.

61E-

071.

84E-21

6.80E-

137.

69E-

153.

68E-

078.26E-

153.

68E-

0710

0ML3/24

134.

58E-

184.31E-

175.28E-

093.

86E-

152.31E-

071.

14E-20

4.84E-

135.23E-

142.37E-

075.

62E-

142.37E-

0710

0ML3/2

514

9.63E-

181.

60E-

178.76E-

091.

13E-

155.

92E-

075.

03E-21

7.52E-

131.43E-

146.

00E-

071.

55E-

146.

00E-

0710

0ML3/2

615

2.02E-

171.

16E-

151.

09E-

084.

89E-

146.

02E-

071.77E-

191.

19E-

127.

87E-

136.

13E-

078.37E-

136.

13E-

0710

0ML3/27

163.

90E-

161.30E-

152.

69E-

087.

04E-

151.

12E-

061.

81E-20

6.40E-

122.46E-

131.

15E-

062.

54E-

131.

15E-

0610

0

Satu

ration Indices

Sam

ple

Solu

tion

si_F

e3(O

H)8

si_A

gmet

alsi

_Boe

hmite

si_F

errih

ydrit

esi

_Hem

atite

si_Cr2

O3

si_D

iaspor

esi

_FeCr2

O4

si_G

ibbs

itesi

_Goe

thite

i_Le

pido

croc

itsi

_Sid

erite

si_Orp

imen

tsi

_Pyr

itesi

_Viv

iani

tesi

_FeS

(ppt

)si

_Rea

lgar

si_C

H4(

g)ML1/2

61

-999

.999

0.14

612.

1791

-999

.999

-999

.999

1.10

073.

8971

-0.0

512.4834

-999

.999

-999

.999

-1.4

8415

.1807

9.2843

-2.8

845

0.31

030.

0811

3.1624

ML1/27

2-9

99.9

990.20

570.72

92-9

99.9

99-9

99.9

990.

8327

2.44

90.

0764

1.03

6-9

99.9

99-9

99.9

99-1

.063

813

.252

89.40

11-1

.2003

-0.0

025

-1.0

997

-1.3

504

ML1/2

83

-999

.999

-0.7

602

0.97

68-9

99.9

99-9

99.9

990.

1 464

2.70

18-0

.756

61.29

12-9

99.9

99-9

99.9

99-1

.1723

13.67

9.4939

-1.2

526

-0.0

66-0

.9754

-1.5

66ML1/2

94

-999

.999

0.6229

0.8028

-999

.999

-999

.999

-0.6754

2.5427

-1.6

851

1.13

89-9

99.9

99-9

99.9

99-1

.4194

14.638

89.

1783

-1.7

020.

0024

-0.3

164

0.6428

ML1/3

05

-999

.999

0.23

940.

6652

-999

.999

-999

.999

-0.7774

2.4122

-1.5

693

1.01

18-9

99.9

99-9

99.9

99-1

.0314

14.5

809

9.672

-0.9

851

0.1726

-0.5

154

-0.6

176

ML3/17

6-0

.953

-2.2

933

0.59

821.

5882

11.038

1.4323

2.3523

0.7928

0.95

944.33

183.

6549

-1.7

199

12.5

5511

.239

9-3

.464

8-0

.184

8-2

.4993

-11.

1365

ML3/1

87

-999

.999

-0.8

129

1.1735

-999

.999

-999

.999

1.69

852.

9258

-0.4

507

1.52

92-9

99.9

99-9

99.9

99-3

.0034

15.4

635

7.89

94-7

.1403

-0.7

896

0.3249

3.6327

ML3/1

98

-1.436

0.42

051.2245

1.83

0711

.5304

1.42

922.

9768

-0.146

11.

5797

4.57

593.

8917

-2.5433

13.1

823

7.5221

-5.8

579

-1.0

011

-0.732

81.

04ML3/2

09

-999

.999

0.99

691.

9324

-999

.999

-999

.999

1.60

693.

6838

0.49

082.28

69-9

99.9

99-9

99.9

99-2

.092

912

.3597

7.8678

-4.9254

-0.7

595

-1.1

951

-0.2

614

ML3/2

110

0.0745

0.3635

1.16

972.23

0912

.3233

1.7144

2.92

120.

8614

1.52

574.

9732

4.28

53-1

.9324

10.578

7.98

02-3

.966

6-0

.9224

-2.223

6-3

.166

9ML3/22

110.2325

1.71

521.

1546

2.13

9212

.140

61.

6097

2.90

611.

0983

1.51

044.

8817

4.1938

-1.6

523

7.1825

6.4446

-3.236

5-1

.7393

-3.5

621

-4.9

88ML3/23

12-9

99.9

992.30

131.

136

-999

.999

-999

.999

2.22

172.

8865

2.01

081.49

03-9

99.9

99-9

99.9

99-1

.426

16.74

046.

1891

-2.542

9-1

.6374

-3.6

033

-4.4332

ML3/24

13-9

99.9

991.

6013

0.77

52-9

99.9

99-9

99.9

991.7642

2.52

581.

8414

1.13

01-9

99.9

99-9

99.9

99-1

.2573

5.53

6.90

88-1

.8723

-1.6237

-4.5

615

-8.4737

ML3/2

514

-999

.999

1.9543

1.287

-999

.999

-999

.999

1.51

63.

0367

0.25

971.

6387

-999

.999

-999

.999

-2.2

065

8.87

976.48

52-5

.5017

-2.1

118

-2.9

179

-4.976

9ML3/2

615

-999

.999

1.54

040.7538

-999

.999

-999

.999

1.52

022.

5034

1.1572

1.10

52-9

99.9

99-9

99.9

99-1

.401

56.

5945

7.3846

-2.7

689

-1.9

854

-4.447

-11.2248

ML3/27

16-9

99.9

992.24

980.

0686

-999

.999

-999

.999

1.4133

1.8174

1.24

650.41

51-9

99.9

99-9

99.9

99-1

.132

86.40

845.

8298

-2.7

851

-2.033

5-3

.7857

-6.1

162

App

endi

x B8

- PH

REE

QC

Out

put

SELECTED X-Y PLOTS TO SUPPORT DISCUSSION IN THE TEXT

Solid Phase Graphs Showing Pearsons Correlations Presented In Table 6.4

Barrier Sands Solid Phase X-Y Plots

As versus Al2O3 in solid phase (barrier sands)

01234567

0 1 2

As versus Ga in solid phase (barrier sands)

01234567

0.0 2.0 4.0

Ga (mg kg-1)

As (m

g kg

-1 )

r2=0.65

Al2O3 (% oxide)

As

(mg

kg-1 )

r2=0.74

As versus Fe2O3 in solid phase (barrier sands)

01234567

0.0 0.2 0.4

Fe2O3 (% oxide)

As

(mg

kg-1 ) r2=0.70

As versus Sb in solid phase (barrier sands)

01234567

0 5 10 15

Sb (mg kg-1)

As (m

g kg

-1 ) r2=-0.06

Appendix B9 – Selected X-Y Plots

Fluvial Sands Solid Phase X-Y Plots

As versus Al2O3 in solid phase (fluvial sands)

0

2

4

6

8

10

12

0 1 2 3

Al2O3 (% oxide)

As

(mg

kg-1 )

As versus CaO in solid phase (fluvial sands)

0

2

4

6

8

10

12

0 2 4

CaO (% oxide)

As (m

g kg

-1 )

r2=0.53r2=0.65

As versus Fe2O3 in solid phase (fluvial sands)

0

2

4

6

8

10

12

0.0 0.5 1.0

Fe2O3 (% oxide)

As (m

g kg

-1 )

As versus Ga in solid phase (fluvial sands)

0

2

4

6

8

10

12

0.0 2.0 4.0 6.0

Ga (mg kg-1)

As (m

g kg

-1 )

r2=0.84r2=0.93

As versus MgO in solid phase (fluvial sands)

0

2

46

8

10

12

0.0 0.2 0.4

MgO (% oxide)

As

(mg

kg-1 )

As versus Sb in solid phase (fluvial sands)

0

2

4

6

8

10

12

0 5 10

Sb (mg kg-1)

As (m

g kg

-1 ) r2=-0.36r2=0.84


As versus SiO2 in solid phase (fluvial sands)

0

2

4

6

8

10

12

70 80 90 100

SiO2 (% oxide)

As

(mg

kg -1

) r2 = -0.73

As versus SO3 in solid phase (fluvial sands)

0

2

4

6

8

10

12

0.0 0.5 1.0

SO3 (% oxide)

As (m

g kg

-1 )

r2=0.78

Fluvial Sand / Estuarine Clay Solid Phase X-Y Plots

As versus Fe2O3 in solid phase (fluvial sands / estuarine clay)

0

2

4

6

8

10

12

0.0 1.0 2.0 3.0

Fe2O3 (% oxide)

As (m

g kg

-1 )

r2=0.01

As versus Sb in solid phase (fluvial sands / estuarine clay)

0

24

6

810

12

0 5

Sb (mg kg-1)

As

(mg

kg-1 )

10

r2=-52


Bedrock Clay Solid Phase X-Y Plots

As versus Al2O3 in solid phase (bedrock clays)

02468

10121416

0 10 20

Al2O3 (% oxide)

As (m

g kg

-1 )As versus Ba in solid phase

(bedrock clays)

02468

10121416

0 500 1000

Ba (mg kg-1)

As

(mg

kg-1 )

r2=-0.93r2=-0.98

As versus Ga in solid phase (bedrock clays)

02468

10121416

0 10 20

Ga (mg kg-1)

As

(mg

kg-1 ) r2=-0.99

As versus K2O in solid phase (bedrock clays)

02468

10121416

0 1 2 3

K2O (% oxide)

As (m

g kg

-1 )

r2=-0.98

As versus LOI in solid phase (bedrock clays)

02468

10121416

0 2 4 6

LOI (%)

As (m

g kg

-1 )

As versus Mo in solid phase (bedrock clays)

02468

10121416

0 1 2 3

Mo (mg kg-1)

As

(mg

kg-1 ) r2=0.60

r2=-0.81


As versus Nb in solid phase (bedrock clays)

02468

10121416

0 5 10 15

Nb (mg kg-1)

As

(mg

kg-1 ) r2=-0.97

As versus P2O5 in solid phase (bedrock clays)

02468

10121416

0.0 0.1 0.2 0.3

P2O5 (% oxide)

As (m

g kg

-1 )

r2=0.90

As versus Rb in solid phase (bedrock clays)

02468

10121416

0 50 100 150

Rb (mg kg-1)

As

(mg

kg-1 ) r2=-0.96

As versus Sb in solid phase (bedrock clays)

02468

10121416

0 2 4

Sb (mg kg-1)

As

(mg

kg-1 )

6

r2=-0.76

As versus SiO2 in solid phase (bedrock clays)

02468

10121416

70 75 80 85 90

SiO2 (% oxide)

As (m

g kg

-1)

r2 =0.95

As versus Sr in solid phase (bedrock clays)

02468

10121416

0 50 100 150

Sr (mg kg-1)

As

(mg

kg-1 )

r2=-0.77


As versus Th in solid phase (bedrock clays)

02468

10121416

0 5 10 15

Th (mg kg-1)

As

(mg

kg-1 )

r2=-0.92

As versus Zr in solid phase (bedrock clays)

02468

10121416

0 100 200 300

Zr (mg kg-1)

As

(mg

kg-1 ) r2=-0.90

Bulk Solid Phase X-Y Plots

As versus FeO in solid phase

02468

10121416

0 5 10

As versus SiO2 in solid phase

02468

10121416

70 80 90 100

SiO2 (% oxide)

As (m

g kg

-1) r2 = -0.54

FeO (% oxide)

As (m

g kg

-1)

r2 = 0.67


APPENDIX C

Photographs

APPENDIX C - PHOTOGRAPHS

Site Photographs

Macleay Estuary

View of Stuarts Point aquifer from the top of Mt Yarrahappinni – looking South

Mt Yarrahappinni behind avocado crops

Mt Yarrahappinni

Arsenic treatment plant

Coffee Rock

Field Work

Arsenic speciation in field

Multi-level piezometers before installation

Close-up of sample point on multi-level piezometer

Installation of multi-level piezometers

Laboratory Work

Preparation of trace element ‘pellets’ for XRF




Sequential extraction procedure

Sequential extraction procedure

SEM

Polished sections for electron microprobe analysis

Bethany Megan O'Shea - UNSWorks - UNSW Sydney

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