TREATING ACIDITY IN COAL PIT LAKES USING SEWAGE AND ...

March 2008 

TREATING ACIDITY IN COAL PIT LAKES USING 

SEWAGE AND GREEN WASTE: MICROCOSM AND 

FIELD SCALE TRIALS AT THE COLLINSVILLE 

COAL PROJECT (QUEENSLAND) 

By, Dr. Clint McCullough 

Dr. Mark Lund 

Mr. Joel May

Prepared for, 

March 2008 

TREATING ACIDITY IN COAL PIT LAKES USING 

SEWAGE AND GREEN WASTE: MICROCOSM AND 

FIELD SCALE TRIALS AT THE COLLINSVILLE 

Centre for Ecosystem Management Report No. 2008-03 

Australian Coal Association Research Program (ACARP) Project: C14052 

Dr. Clint McCullough BSc, MSc (Hons), PhD (Aquatic Ecotoxicologist), 

Dr. Mark Lund BSc (Hons), PhD (Aquatic Ecologist), 

Mr. Joel May BE (Environmental Engineer). 

COAL PROJECT (QUEENSLAND) 

1 

By, Dr. Clint McCullough 

Dr. Mark Lund 

Mr. Joel May

Treating acidity in pit lakes using sewage and greenwaste 

1 Executive Summary 

• This research project involved, collaboration and technology transfer from research in 

Collie (Western Australia) to Collinsville (Queensland). 

• Coal mine lakes represent a potentially valuable resource to both the environment and 

the community in inland Australia, if the water can be treated (or remediated) to an 

appropriate standard for its proposed end use. Beneficial end uses include: 

aquaculture, water for irrigation, recreation, and for nature conservation. Research in 

Collie coal mine lakes and internationally has found there is considerable potential for 

utilising biological processes to reverse acidity generating processes. This approach 

focuses primarily on additions of organic material to support sulfate reducing bacteria 

(SRB) which convert sulfate back to sulfides, removing acidity and metals in the 

process. The approach also fosters a range of other biological processes which can 

increase alkalinity and pH. 

• Previous ACARP funded research in Collie had tested the use of SRBs for treating 

acidic mine waters with limited success. Limiting the effectiveness of this approach in 

Collie are low sulfate concentrations that occur in these pit lakes. Nevertheless, there 

are a number of highly acidic mine lakes in the Collinsville Coal Project (CCP) with 

high sulfate levels. 

• CCP discharged 4 ML of highly acidic mine water (typical of their pit lakes) into a 80 

ML sewage evaporation pond. CCP and later the authors (CM & ML) monitored the 

effects of this discharge on water quality in the ponds. In approximately 18 months 

water quality in the evaporation pond had returned to pre-addition conditions. It 

appeared that a combination of processes, including SRB activity were responsible for 

the remediation of the mine water. The results of this study suggested that SRB 

activity might be useful in the treatment of pit lakes on the CCP site. 

• This project aims to establish a full scale demonstration of the application of passive 

biological remediation in a pit lake on the CCP lease. This pit lake is located near to 

the Bowen Shire’s Collinsville Water Treatment Plant (CWTP) and green waste 

transfer station. Organic substrate in the form of primary treated sludge from the plant 

were be added to the lake. This wastewater provided a ready source of, BOD 

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(Biochemical Oxygen Demand, to reduce oxygen levels to those suitable for SRBs), 

available carbon (C), and nutrients (N & P). All of these wastewater components have 

been identified as key factors promoting SRB activity. Monitoring of this 

experimental lake and control lakes both before and after the addition enabled an 

assessment to be made of the success of this innovative remediation method. 

• The first two experiments using acrylic tubes (microcosms) containing sediment and 

pit lake water were established firstly in Collinsville and then in Perth. These 

experiments were used to gain a better understanding of the processes responsible for 

mine water remediation and to provide estimates of the quantities of organic matter 

that were likely to be required in a full-scale treatment of a pit lake. 

• The first microcosm experiment involved testing of greenwaste only, sewage only, 

and greenwaste and sewage for their effectiveness in remediation of pit lake water, in 

typical Collinsville conditions. 

• This first experiment clearly showed remediation of pit lake water pH from 2.2 to 5.5 

in 145 days, with commensurate declines in iron, aluminium and toxic metal 

concentrations. The presence of greenwaste appeared to be important for the 

effectiveness of the treatment. This was mainly believed to be due to the bulk of the 

greenwaste which extended the area of SRB activity up through the water column. 

• The second experiment repeated the first experiment in design but added different 

quantities of organic material and a second sewage type. Another important aim of 

this experiment was to demonstrate the repeatability of the initial results given the 

heterogeneous nature of the organic materials being used. 

• This second experiment demonstrated that results were generally reproducible from 

the first experiment. However the performance of greenwaste only was poorer with a 

pH of >4 reached after 210 days, while the other treatments produced circum-neutral 

pH (~7) after 120 days with the exception being low levels of sewage only which 

produced a pH of ~6 after 210 days. Associated with the pH improvements were 

reductions in electrical conductivity, sulfate and metal concentrations. Key drivers for 

rapid remediation appeared too be the high tropical temperatures, the combination of 

greenwaste and sewage, and a ‘fresher’ type of sewage from Bowen rather than 

Collinsville. This experiment enabled the team to determine the quantities of organic 

materials required for a field trial. 

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• Comprehensive risk assessments and stakeholder consultation identified significant 

occupational health and safety concern of using sewage and greenwaste at field scale 

was of high levels of faecal coliforms in the water creating a risk for operators. The 

second experiment revealed that after 180 days there were no faecal coliforms in the 

water, although total coliforms were high. These latter coliforms were believed to be 

associated with decomposition process and were not believed to pose a contact risk to 

operators. Nevertheless, precautions were developed for operators involved in the 

project. 

• The field scale trial followed 15 months of water quality monitoring on three CCP pit 

lakes. These data formed part of the before-after-control impact (BACI) design of the 

field experiment. 

• A review of the data from the second experiment, indicated that given material 

availability that a smaller pit lake was need for treatment. In July 2006, it was decided 

to split with an earth wall the Garrick Area East pit lake into a smaller 71 ML 

(GAEW) for treatment and the remaining 336 ML (GAEE) to be kept as an untreated 

control. 

• In August 2006 to January 2007, the smaller lake (GAEW) was filled with dried 

sewage sludge (60 t), liquid sewage sludge (3,190 t) and municipal green waste 

(980 t). Monitoring of this new treatment lake and the remaining control lake and 

other control lakes then continued for another 6 months at monthly intervals. 

• Due to groundwater influx and heavy cyclonic rainfall events, it was often unclear 

what contribution sulfate reduction process have made to changes seen in water 

quality. Nevertheless, physico-chemical changes to control lakes during monitoring 

could generally be explained as a result of these two external influences. However, 

after four months of filling ceasing, GAEW ORP began to decline from around 

600 mV to 200 mV starting from the benthos. Also beginning at the benthos, pH 

increased soon afterwards reaching a pH of 3.7 across the lowest 3 m of water column 

by July 2007. The lower 3 m mean pH of the GAEE lake was 2.2 at this time. 

Similarly, at the end of the experiment electrical conductivity was reduced to 

9.0 mS cm -1 compared to 9.4 mS cm -1 in the GAEE lake. 

• These field chemistry observations suggest that addition of low-grade organic 

materials for remediation of acid mine waters at field scale shows promise. However, 

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remediation has only just begun and further monitoring is required to access the 

degree of treatment that can be achieved and how long this treatment will continue. 

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Frontispiece 

Figure 1. Joel May taking a post-organic dosing water sample from treatment lake GAEW. 

This document should be referenced as follows. 

McCullough, C. D.; Lund, M. A. & May, J. M. (2008). Treating acidity in coal pit lakes 

using sewage and green waste: microcosm and field scale trials at the 

Collinsville Coal Project (Queensland). ACARP Project Report C14052. 

Centre for Ecosystem Management Report p2008-03, Edith Cowan 

University, Perth, Australia. 112pp. Unpublished report to Australian Coal 

Association Research Programme. 

7

2 Contents 

9 

McCullough, Lund and May (2008) 

1 Executive Summary 3 

2 Contents 9 

3 Introduction 11 

3.1 Aims 13 

4 The Collinsville Coal Project Area 15 

5 Effects of adding mine lake water to a sewage evaporation pond 19 

5.1 Overview 19 

6 Collinsville microcosms - Experiment 1 21 

6.1 Background 21 

6.2 Methods 24 

6.3 Results 26 

6.4 Discussion 34 

7 Perth microcosms – Experiment 2 37 



7.2.1 Statistical analyses 43 


7.3.1 Changes in water quality by Day 180 of the experiment 47 

7.3.2 Sediment 50 

7.3.3 Changes in water quality during the experiment 53 

7.3.4 Bacterial abundances 55 

7.3.5 Gases evolved 56 


8 Field-scale remediation of a tropical acid mine pit lake with 

greenwaste and sewage sludge 61 



8.2.1 Risk Assessment 65 

8.2.2 Organic dosing 66


8.2.3 Sampling 68 

8.2.4 Groundwater interactions 68 

8.2.5 Pit lakes 69 


8.3.1 Climate 69 

8.3.2 Dosing with organic material 70 

8.3.3 Physico-chemical parameters 72 

8.3.4 Metal/metalloid concentrations 79 

8.3.5 Nutrients 87 

8.3.6 Groundwater interactions 90 


9 Conclusions 99 

10 Acknowledgements 102 

11 References 103 

12 Appendix: Publications and Presentations arising from ACARP 

Project C14052 113 

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3 Introduction 

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Environmental legacies of open cut mining operations are the large voids created. Too large 

to backfill and typically requiring extensive dewatering operations to keep dry, many are 

destined to become large, deep lakes. In Australia, these new lakes have few natural 

counterparts in size (especially depth) and are a potentially valuable new asset to rural 

Australia. Despite considerable improvements in mining practices, acid mine drainage 

(AMD) resulting from exposed overburden and void walls can result in these new lakes 

becoming highly acidic. Worldwide AMD is potentially the largest negative environmental 

impact resulting from coal mining (Robertson, 1987; Ryan & Joyce, 1991; Lowson et al., 

1993; Harries, 1998a, b). In Australia there are many well publicised examples of the 

consequences of AMD from coal and other mineral mines including Rum Jungle (Northern 

Territory), Captains Flat (New South Wales), Brukunga (South Australia), Mount Lyell 

(Tasmania), and Mount Morgan (Queensland) (Zhou, 1994; Lawton, 1996). 

Extremely elevated aluminium, iron and sulfate concentrations, combined with extremely low 

dissolved carbon and other nutrients, are typical of AMD (Castro et al., 1999; Peiffer et al., 

1999; Fyson, 2000; Woelfl, 2000; Taylor & Waters, 2003). Dissolved metal concentrations 

may be so high as to be toxic to most aquatic biota (Spry & Wiener, 1991; Bowell, 2000). 

Although literature is sparse for sulfate toxicity (ANZECC/ARMCANZ, 2000a), 

concentrations can also reach potentially toxic levels in these lakes. However, acid mine 

drainage (AMD) waters also differ greatly between regions and geologies (Robb & Robinson, 

1995; Johnson & Hallberg, 2003). 

Australian data on remediation techniques specific to our pit lakes is sparse. The expensive 

and high-maintenance operations of active remediation strategies (e.g. liming, anoxic 

limestone drains) are often not appropriate to the remoteness, high dissolved metal 

concentrations and high acidity of many Australian mine lakes (Brown et al., 2003; Taylor & 

Waters, 2003). Furthermore, although dealing with low pH issues of AMD waters, active 

remediation strategies are often not efficient at removing high dissolved concentrations of 

metals such as iron and aluminium, or sulfate.


In natural systems and in old mine lakes which have accumulated organic materials in the 

sediments, sulfate is removed by SRB reduction in carbon-rich and anaerobic conditions 

producing hydrogen sulfide (Frömmichen et al., 2003). Hydrogen sulfide is chemically 

highly reactive and binds strongly to heavy metals such as iron, copper, cadmium and zinc 

forming insoluble precipitates (Kleeberg, 2000; Brown et al., 2003). Furthermore, elevated 

metal ions may also be removed by direct sorption and complexation to organic matter 

(Peiffer et al., 1999; Fyson, 2000). 

Controlled organic enrichment (saprobisation) of lake sediments in laboratory microcosm 

experiments has been found to increase pH and decrease acidity as a result of SRB processes 

(Castro et al., 1999; Frömmichen et al., 2003). Sulfate and iron reducing bacteria are a 

diverse and ubiquitous group of micro-organisms (Johnson & Hallberg, 2003). However, 

their activity is directly limited by the low carbon availability typical of waters in acid pit 

lakes (Kafper, 1998). In addition, to sufficient available carbon SRB require an anoxic (no 

oxygen) environment to reduce sulfate. In many pit lakes, which stratify over summer, the 

hypolimnion (bottom waters) can take a considerable time to become anoxic due to low 

biological oxygen demand (BOD) in the sediments (Boland & Padovan, 2002). Addition of 

organic matter can provide available carbon and high BOD providing suitable conditions for 

SRB activity (Klapper et al., 1998). Most of the research into organic matter additions has 

come from the northern hemisphere and has focused on expensive and largely unavailable 

sources of carbon for an Australian mine site; such as glucose, ethanol, molasses, whey, 

potato processing waste and cattle effluent (Castro et al., 1999; Frömmichen et al., 2003). 

Nonetheless, other than establishing that there is a requirement for an easily accessible supply 

of carbon, little is understood about differences in the quality of organic substrate for 

bacterially-mediated sulfate reduction (Blodau et al., 2000). 

In Collie (Western Australia) pit lakes, experiments have trialled a number of organic matter 

types in experiments at scales ranging from laboratory, to ponds to mesocosms in pit lakes 

(Lund et al., 2000; Phillips et al., 2000; Lund, 2001). Readily available organic materials 

such as mulch, hay, manure (cow) and sawdust have been tested (Thompson, 2000). 

Although there have been noticeable improvements in biodiversity within the experiments 

with organic matter additions, there has been little long term change in pH. This has been 

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attributed to low sulfate concentrations (typically


The project team undertook the following projects: 

• Assessment of the capacity for a sewage evaporation pond to treat acidic pit lake 

water. 

• Assessments in microcosms of the efficacy of sewage and greenwaste to treat acidic 

pit lake waters in Collinsville. 

• Field scale trial of the use of sewage and greenwaste to treat a pit lake in Collinsville. 

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4 The Collinsville Coal Project Area 

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Collinsville is a typical small inland Australian mining town located approximately 70 km 

from the coast of North Queensland, Australia (Figure 4.1). Together with the mining town of 

Scottville, Collinsville is located within a kilometre of the active mining Collinsville Coal 

Project coal mining lease (Figure 4.2). Collinsville was underground (bord and pillar) mined 

for its coal as early as 1919 at the State Mine working the Bowen Seam, and began open-cast 

mining in the mid 1950s. Mining operations are expected to continue until at least 2013. The 

combined population of Collinsville and nearby Scottville in 2006 was 2,664 people and is 

undergoing growth as mining activity increases. 

Figure 4.1 Location of the study site in Collinsville, North Queensland, Australia.


Figure 4.2. CCP Project Lease pit lakes and surrounding towns. 

Collinsville has a semi-arid tropical climate with a rainfall regime that falls into a transition 

between sub-humid and semi-arid. The climate is dominated by a moderately low, highly 

episodic and unreliable summer rainfall (708 mm/annum) and a very high evaporation rate 

(1,860 mm/annum) (Commonwealth of Australia Bureau of Meteorology, 09/02/2005; 

Davies & Willcocks, 1992) (Figure 4.3). 

Figure 4.3. Mean temperature and rainfall climate of Collinsville (Commonwealth of Australia Bureau of 

Meteorology, 09/02/2005). 

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The geology of the region is of highly weathered hard rocks with soils of very low organic 

carbon content. Surrounding vegetation is predominantly Eucalyptus and Acacia spp. 

dominated open woodland with an annual grass understory. There are some 20 pit lakes in the 

CCP lease, all of which are acidic with high concentrations of dissolved solutes. All of these 

lakes are of extremely low pH (ca. pH 2) and contain high concentrations of dissolved solutes 

(electrical conductivity = 9–19 mS cm -1 ). One of these acid pit lakes, Garrick Area East 

(GAE), is the focus of this study due to its proximity to the Collinsville waste water treatment 

plant and green waste dump (


Figure 4.5. Garrick Area East pit lake at CCP. Note very dark orange colour of water due to high dissolved 

iron concentrations (pH = ca.2, dissolved iron concentration = 1,100 mg L -1, electrical 

conductivity = ca.9 mS cm -1 ), and efflorescence of elemental sulfur, gypsum and epsomite around 

littoral fringe. 

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5 Effects of adding mine lake water to a sewage 

evaporation pond 

The project team has been involved in another project at Collinsville that, while outside this 

ACARP project, contributed to its development. Unfortunately as this work has been 

published elsewhere it cannot be included here. We have included an overview to illustrate 

why we had confidence that the use of sewage and greenwaste was likely to be successful. 

5.1 Overview 

The addition of acidic mine water to raw sewage and workshop wastewaters in an 

evaporation pond provided an opportunity for a field-scale experiment as essentially a 

reversal of in-situ treatment of acidic pit lakes by addition of organic carbon. The hyper- 

eutrophic evaporation pond initially contained high concentrations of nutrients, a pH >8, high 

levels of sulfate (500 mg L -1), and had regular algal blooms. Soon after the addition of the 

AMD pit water, the evaporation pond pH fell to 2.4, and electrical conductivity (EC) and 

most metal concentrations were elevated by one to two orders of magnitude. Over the 

following 18 months, the pH of the pond increased and the EC and metal concentrations 

decreased. After only 18 months of addition of AMD, pond water quality had returned to a 

level similar to that before AMD addition. These observations suggest that addition of low- 

grade organic materials shows promise for remediation of acid mine waters at field scale and 

warrants experimental investigation. 

McCullough, C. D.; Lund, M. A. & May, J. (in press). Field scale demonstration of the 

potential for sewage to remediate acidic mine waters. Mine Water and 

Environment 27(1). DOI: 10.1007/s10230-007-0028-y

6 Collinsville microcosms - Experiment 1 

A version of this chapter has been published without copyright restrictions as: 

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McCullough, C. D.; Lund, M. A. & May, J. M. (2006). Microcosm testing of municipal 

sewage and green waste for full-scale remediation of an acid coal pit lake, in 

semi-arid tropical Australia. Proceedings of the 7th International Conference on 

Acid Rock Drainage (ICARD). St Louis, Massachusetts, USA. 1177-1197. 

6.1 Background 

Microbial sulfate reduction is considered to be an efficient and effective remediation for the 

treatment of acid mine drainage (AMD), through alkalinity production and precipitation of 

metals as sulfides through reaction with sulfide cations (Equation 1) and as carbonates 

through increased pH (Equation 2) (Dillon et al., 1997; Küsel & Dorsch, 2000; Benner et al., 

2002; Praharaj & Fortin, 2004). In-lake neutralisation via sulfate reduction is expected to play 

a key-role in the remediation of acidic mining pit lakes (Kleeberg, 1998). However, since 

Tuttle et al. (1969) first suggested the use of sulfate reducing bacteria (SRB) in the treatment 

of AMD, treatment has largely focused on ex-situ treatment in bioreactors. What in-situ 

treatment relevant experiments that have occurred have generally been performed at 

microcosm (e.g., carboy vessels; (Fyson et al., 1998; Castro et al., 1999; Frömmichen et al., 

2004)) or at best macrocosm (e.g., limnocorrals) scales (Martin et al., 2003). Only recently 

has attention switched to in-situ remediation systems at actual field treatment scales (Gibert 

et al., 2002). 

Equation 1 Metal 2+ + SO4 2- + 2C(organic) → MetalS + 2CO2 

Equation 2 Metal 2+ + CO3 2- → MetalCO3 

A review by Gibert et al. (2002) found that the nature of the organic matter was a prime 

determinant of the efficacy of the passive treatment system. For example, the availability of


carbon from plant matter is dependent upon decomposition, which is extremely limited in 

acidic and anoxic conditions (Harris & Ragusa, 2001). To this end, Waybrant et al. (1998) 

found mixtures containing multiple sources of organic matter demonstrated higher sulfate 

reduction rates than those of single sources. Castro and Moore (1997) observed that, 

“Therefore, for economic feasibility the added organic matter must be cheap and locally 

available.” Hard et al. (2003) echoed this view noting that, “For a microbial process to be 

economically feasible, the carbon and energy source should be cheap, widely available and 

highly effective.” However, most experiments into the utility of sulfate reduction processes 

for ameliorating AMD have instead focused upon highly labile but expensive carbon 

substrates such as ethanol (Kolmert & Johnson, 2001; Martin et al., 2003; McNee et al., 

2003), sugar (Pöhler et al., 2002; Wendt-Potthoff et al., 2002; Frömmichen et al., 2003; 

Geller et al., 2003; Frömmichen et al., 2004), cow manure (Drury, 1999, 2000), etc. For 

many remote mining locations, these materials are unviable for practical and economic 

reasons. The only bulk carbon sources likely to be available in these areas will be municipal 

sewage, and green waste (including a broad range of plant material collected by local 

government from domestic and municipal lawns and gardens). 

Using readily-available and economically-viable sources of organic materials, Waybrant et al. 

(1998) found all of the eight organic matter types they tested reduced sulfate, with sewage 

sludge the fastest to achieve high levels of sulfate reduction. The mixture of sewage sludge 

and green waste (leaf mulch, woodchips and sawdust) reduced 4,500 mg L -1 of sulfate to 

3 within 30 days of SRB activation and simultaneously decreased 

divalent metal concentrations. This mixture proved more effective in ameliorating pH and 

metal concentrations than either sewage sludge (little response) or plant material (nil 

response). Gusek (2002) even suggested injecting sewage sludge into mine shafts and adits to 

remove and prevent production of acidity from these sources. 

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More specific data for our particular study area was reported in a 1993 study at the 

Collinsville Coal Project Mine in North Queensland, Australia (Fallon, 1994). Although 

unquantified, this report details findings that the extremely acid waters of Blake A-cut had 

“notably improved” (i.e. reduced acidity, sulfate and metal concentrations) and suggested that 

one mechanism may be bacterially-mediated sulfate reduction occurring with the sewage 

effluent discharged into this pit being used as a source of carbon. As rates of SRB activity are 

enhanced at warmer temperatures, we expect that the warmer tropical conditions of 

Collinsville will enhance activity especially compared to that described in the majority of 

published literature which is from the cooler temperate regions of Europe and North America. 

Consequently, there is a large body of published data to suggest that bacterially-mediated 

sulfate reduction processes can ameliorate acid mine drainage waters. Published literature 

from laboratory and mesocosm experiments also indicates that sewage is suitable for use as 

an organic substrate stimulating sulfate reduction. However, there are few published reports 

of field scale attempts of AMD lake remediation. Furthermore, very little bioremediation 

work has been carried out on the mining pit lakes of Australia using these methods, especially 

in semi-arid tropical Australia where a significant number of acid pit lakes historically occur 

and are still being developed of increasing sizes (Harries, 1997). As an arid continent with 

increasing pressure on water resources, there is growing demand for new sources of water to 

meet a variety of end uses (Doupé & Lymbery, 2005). 

The end use of many of these pit lakes may also often be only for slightly remediated water 

quality of lower salinity for mining operation dust suppression or similar industrial use. 

Consequently, financially viable treatment to this lower standard may still be very achievable 

in even remote mining areas. Consequently, this first experiment was intended to test the 

potential of using sewage and greenwaste in concert or individually to remediate GAE pit 

lake water.


6.2 Methods 

Microcosm experiments were designed to mimic the hypolimnetic water column and 

sediment regions of a typical strongly stratified Collinsville pit lake. Twelve clean 100 mm 

diameter and 600 mm long (4.5 L) acrylic tubes were set up in an uninsulated laboratory on 

the mine site as microcosms containing 140 mm of sediment and 440 mm of GAE pit lake 

water. Cores were pushed into the littoral sediment of GAE to a depth of 200 mm and sealed 

with rubber bungs at the top and bottom. The height of sediment in the bottom of the core 

was then adjusted by sliding it downwards out the bottom of the core to a final depth of 

140 mm across all cores. Three replicate cores were then allocated to each of the following; 

control (untreated), green waste only (G), sewage (S), or green waste and sewage (GS) at 

dosing rates realistic of that able to be achieved in a typical CCP pit lake using regional 

sources for several month’s filling. 

Green waste was sourced from the Collinsville shire green waste dump nearby the 

Collinsville Coal Project lease. Greenwaste consisted of a wide range of garden clippings 

from both woody and herbaceous species that had been exposed to Dry season drying 

climatic conditions for some weeks. Primary-treated sewage sludge was sourced from the 

Collinsville Municipal Wastewater Treatment Plant. Due to the over-capacity of the plant for 

the declining town size, the sewage sludge had been exposed to the sun in drying beds for 

around 12 weeks. 

The cores were then filled with GAE water to within 100 mm of their brim, green waste was 

then added where appropriate, and then sewage in the following scheme (Table 6.1). 

Microcosm core water levels were topped-up to with 20 mm of the core’s brim and a loose- 

fitting rubber bung was applied to the top (not airtight) to reduce air infiltration into the core 

water as would a strongly stratified epilimnion (Figure 6.1). The entire suite of microcosm 

cores were then placed in a 500 mm high opaque black plastic planter tub which was filled 

with water to evenly distribute ambient temperatures between cores and to reduce the 

incidence of leakage. The lower water-filled tub was then capped with an identical inverted 

planter tub to occlude light, as PAR levels are extremely low in the hypolimnion of these 

lakes (author’s unpublished data). 

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The cores were then sampled for the physico-chemical variables; temperature, pH specific 

conductance, oxidation-reduction potential and dissolved oxygen (% saturation and mg/L) 

with a Hydrolab Quanta multiparameter meter. 

Table 6.1. Experimental design for organic dosing of pit lake cores. 

Treatment level Organic dosing mass Water:green Number of 

(g) 

waste:sewage ratio replicates 

Control 0 1:0:0 3 

Green waste 200 16:1:0 3 

Green waste and sewage 100 and 200 32:1:2 3 

Sewage 200 16:0:1 3 

Figure 6.1. Microcosm cores on day 0. Treatment allocations from left to right are control, green waste, 

sewage and green waste and sewage. 

Physico-chemical measurements were taken at one day after filling, two days after, and then 

at approximately weekly intervals thereafter for 145 days. 

A 100 mL water sample was taken after 82 days and analysed for concentrations of solutes. 

Cations (aluminium, arsenic, bromine, calcium, cadmium, chromium, cobalt, copper, iron, 

lead, magnesium, mercury, sulfate, selenium, silica, tin, zinc) were analysed via ICP-AES.


Ammonia and NOx (nitrate and nitrite) anions were analysed on an auto-analyser using the 

Berthelot and persulfate digestion methods respectively, with subtraction (APHA, 1998). 

Primary treated municipal sewage was collected from drying beds of the Collinsville Water 

Treatment Plant. Due to a very low output from the plant, this material had been dried for 

around 12 weeks and was thus of an extremely low water content (specific 

density = 1.35 kg/L). A sample from this sewage was taken for analysis. 

Green waste (specific density = 1.2 kg/L) was collected from the Collinsville green waste tip. 

This material was largely representative of Australian garden waste and included lawn 

clippings, palm fronds and other leafy and woody material. Green waste was chopped into 

approximately 50 mm sections to fit into the cores. A representative sample of green waste 

was prepared in Perth (where the chemical analysis was conducted) from similar local plants 

and was dried at 25 o C for one week to simulate ambient conditions at Collinsville municipal 

green waste dump. The sewage and green waste samples were then analysed for total 

nitrogen, total phosphorus and total organic carbon content as well as for dissolved heavy 

metals 

6.3 Results 

Garrick East pit lake had a very low pH and high total iron concentrations indicative of an 

iron buffered system (Table 6.2), with heavy metals at environmentally toxic concentrations 

(ANZECC/ARMCANZ, 2000b). Nutrient levels were also very low in this pit lake; 

especially for phosphorus as is typical of AMD waters (Lessmann et al., 2000; Borg & Holm, 

2001). 

Collinsville sewage sludge was high in sulfur (10 g kg -1 ) and also contained notable levels of 

aluminium (17 g kg -1 ), iron, calcium and magnesium. However, concentrations of some 

heavy metals including zinc (1,500 mg kg -1 ) and nickel (39 mg kg -1 ) were also surprisingly 

high. S was moderately high in nitrogen (31 mg kg -1 ) and phosphorus (12 mg kg -1 ), had 

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relatively high organic carbon content (29%) and contributed significant alkalinity to the 

dosed AMD water as indicated by the alkaline pH paste test (Table 6.2). 

Although G contributed less alkalinity to the AMD water, it contained the highest percentage 

of organic carbon (39%) (Table 6.2). G also contained moderate amounts of sulfur (1,700 mg 

kg -1), nitrogen (20 mg kg -1) and phosphorus (2.3 mg kg -1). Although calcium and magnesium 

levels were higher in G compared to S, heavy metal concentrations were very low. 

Table 6.2. Chemistry of GAE pit lake water and organic materials used in core experiments. 

Parameter GAE water 

(mg L -1 ) 

Sewage 

(mg kg -1 ) 

Total sulfur No data 10,000 1,700 

Green waste* 

(mg kg -1 ) 

Sulfate 2,610 No data No data 

Total nitrogen 0.51 П 31 П 20 

Total phosphorus


Although the control treatment looked largely unchanged over the course of the experiment, 

G had darkened considerably; probably through leaching of labile organic material and with a 

sweet odour indicative of fermentation processes. The presence of this process was further 

reinforced by the development of a blue-green coloured mould on the surface of both G and 

on GS throughout the experiment (Figure 6.2). S became more orange in colour, although 

when inspected from above this was found to be due to a precipitate (probably iron) settling 

on the inside of the acrylic tube. The water in S appeared to have lost the orange tinge which 

the control cores still retained. The space between the GS and the sediment surface began to 

blacken in GS after only a few weeks of the experiment beginning. After Castro et al. (1999), 

this black material was assumed to be ferrous sulfides from the activities of sulfate reducing 

bacteria, as the space above the mesocosm sediments of this treatment which were likely to 

be anaerobic. A strong sulfide smell, (confirmed as hydrogen sulfide using an ITX- Industrial 

Scientific Multigas Meter), also evolved as the black precipitate extended upwards through 

the mesocosm above GS solids. This smell and the presence of black precipitate indicated 

that sulfate reduction was occurring in these GS treatments. 

Two days after addition of organic substrates to AMD cores there was a statistically 

significant mean increase in pH above the control for the treatments of GS, and S 

(F 3,8 = 23.060, p0.05). All treatments also showed a significant decrease in redox 

compared to the control (F 3,8 = 16.673, p = 0.001). 

Figure 6.2. Mould growing on the surface of green waste-only cores at day 72. 

28

29 


Figure 6.3. Microcosm core mean pH and redox potential on day 2 after addition of organic substrates. Error 

bars indicate single standard errors of the mean. Bars with same letter are not significantly 

different, bars with different letter are significantly different (p


(i.e., day 37). Although treatments containing sewage were very variable between replicates, 

both treatments containing sewage appeared to decline in solute concentrations greater than 

did green waste alone. 

Figure 6.4. Change in a). pH, and b). specific conductance, of core microcosm treatment levels over time. 

Error bars indicate single standard errors of the mean. 

30 

a). 

b).

31 


Aside from iron and magnesium which increased in the G treatment; and cadmium, calcium, 

copper, lead and zinc which increased in the S treatment, other solutes had decreased in all 

treatments by day 86 (Figure 6.5). However the only statistically significant difference 

(P


Figure 6.5. Day 86 concentrations of a). major, and b). minor, analytes for each treatment level. Error bars 

indicate single standard errors of the mean. 

32 

a). 

b).

33 


A Principal Components Analysis (PCA) indicated that addition of organic matter appeared 

to increase the variability of water chemistry parameters over that of the control (Figure 6.6). 

There also appeared to be synergistic effects on variability with the combined GS treatment 

producing the most varied response between replicates. Addition of green waste moved the 

water chemistry primarily along the first principal component axis, addition of sewage 

separated along the second principal component axis. After 86 days, addition of sewage was 

correlated with increases in dissolved concentrations of the heavy metals cadmium, copper, 

lead and zinc; and ammonia. Addition of sewage was correlated with decreases in iron, 

chromium, sulfate and electrical conductivity. Addition of green waste was correlated with 

increases in pH and with decreases in dissolved oxygen, redox, aluminium, nickel, sulfate 

and selenium. Nevertheless, although more variable, water chemistry correlations with green 

waste and sewage were very similar to green waste. 

Figure 6.6. PCA of day 86 microcosm core solute concentrations and physico-chemistry. Vectors indicate 

variables contributing │>0.25│ to an eigenvector. Ellipses represent range of variation within 

treatment levels.


6.4 Discussion 

The decrease in sulfate concentrations in all of the treatment microcosms provides further 

support for biological sulfate reduction occurring at varying rates and removing cations as 

precipitates in the dosed mesocosms. However, the success of sulfate reduction in low pH 

waters appears to be contrary to Postgate’s (1984) assertion that sulfate reducing bacteria 

(SRB) activities only occur at pH >5. Nevertheless, published studies some time ago have 

reported SRB activity at pH values below this threshold. For example, SRB activity in an 

acid lake has been found by Herhily and Mills (1985) at pH 2.5, and at pH 2.7 by Gyure et al. 

(1987). In these studies and our own, it appears that slightly-acid microclimates, facilitated in 

part by alkalinity produced through some sulfate reduction, enabled sulfate reducing bacteria 

to survive and to chemically reduce sulfate, producing alkalinity as part of this biological 

process (Küsel & Dorsch, 2000; Küsel et al., 2001). 

Depending on its source, sewage sludge can contain high concentrations of heavy metals 

(Berrow & Webber, 1971). This is especially so of cadmium, which is often concentrated in 

sewage sludge from vegetables which have been fertilised with inorganic phosphates; a 

common application in depauperate Australian soils (Nursita et al., in press). Unexpectedly 

for a non-industrial content sewage source, the Collinsville sewage also displayed high 

concentrations for nickel, lead, and zinc. However, these high heavy metal concentrations are 

still unlikely to present a problem over a longer remediation time. Along with iron, these 

toxic heavy metals are expected to precipitate out of solution by reaction with hydrogen 

sulfide, such that their toxic effect is removed from the water column to levels lower than 

observed at day 86. Further reduction in biological availability is likely to occur through 

formation of complexes with organic chelators present as components of the refractory green 

waste, such as organic acids (Tipping & Hurley, 1992). Nevertheless, the contribution that 

organic materials may make to the overall heavy metal burden of a pit lake needs to be 

considered in any remediation project. This concentration of heavy metals presents an 

opportunity for minerals reprocessing, one which other mining companies (including Xstrata 

PLC. LTD. involved in this project) are investigating (Zinck, 2006). This study found a 

straight green waste treatment performed similarly to that of green waste and sewage, with 

fewer heavy metals contributed to the water column. In this respect, remediation strategies 

34

35 


may be best placed by choosing green waste as the bulk contribution to electron donors over 

that of sewage. As discussed, green waste also has an additional advantage of providing 

organic substances such as humic and fulvic acids, with which heavy metals may directly 

complex to. Ligand formation between heavy metals and refractory organics will further 

remove these toxic components from biological availability, albeit at a likely reduced 

capacity to that of sulfate reduction processes (Brown Jr., 2001). 

The initiation of alkalinity production in green waste and green waste and sewage in only 

37 days is likely to be due to a combination of two important factors unique to this study. 

Firstly, most other research published to date has occurred in cool temperate areas of Europe 

and North America e.g., (Küsel & Dorsch, 2000; Küsel et al., 2001; Benner et al., 2002; 

Tostche et al., 2003; Frömmichen et al., 2004). The higher temperatures experienced in 

Collinsville, even during the middle of the Dry season (19.5–28.2 o C) are likely to have 

exponentially increased biochemical rates of carbon diagenesis and metabolism. Seasonal 

change in ambient temperatures has also been identified as limiting rates of sulfate reduction 

in some experiments with reduced rates of sulfate reduction occurring over cooler seasonal 

periods (Gammons et al., 2000; Benner et al., 2002). 

Secondly, the use of largely fresh green waste as opposed to refractory organic substrates 

such as straw (Frömmichen et al., 2003; Frömmichen et al., 2004), rye grass (Harris & 

Ragusa, 2000, 2001), etc., distinguishes this research from many others in the published 

literature. The greater labile fraction of organic material available in this fresher material may 

have contributed directly to electron donors for sulfate reduction. For example, chlorophyll 

was seen to be leached from the green waste and it is likely that highly labile sap sugars 

would have leached also. Consequently, the use of fresh green waste to acid pit lakes, with or 

without complementary additions of sewage, may prove to be a novel practicable remediation 

strategy for AMD issues in remote mining locations.

7 Perth microcosms – Experiment 2 


37 


Open-cut mining can create pits that are below the natural watertable. Once dewatering 

operations stop, these pits will form pit lakes as surface and groundwaters equilibrate (Castro 

& Moore, 1997). Mining can lead to the exposure of rock strata to weathering which can 

result in acidification and metal contamination of contacting waters as in Acid Mine Drainage 

(AMD) (Banks et al., 1997). Receiving environments for AMD typically have reduced 

environmental and social values, and the resultant water is less valuable as a resource to the 

mining company (McCullough & Lund, 2006). Acid Mine Drainage is arguably the greatest 

environmental problem facing water management in the international mining industry today 

(Gray, 1997; Harries, 1998a). The large quantities of water in pit lakes potentially represents 

a potentially valuable resource to mining companies, the environment and community; if 

appropriate water quality can be achieved (Doupé & Lymbery, 2005; McCullough & Lund, 

2006). 

Sulfate reducing bacteria (SRB) can reverse the acidification process by converting sulfates 

to sulfides in low redox environments when supplied with labile carbon sources (see King et 

al, 1974). Dissolved metals can bind with sulfides or ultimately carbonates to form insoluble 

precipitates increasing pH.(Dillon et al., 1997; Hard et al., 1999; Küsel & Dorsch, 2000; 

Benner et al., 2002; Praharaj & Fortin, 2004). Tuttle et al. (1969) first suggested the use of 

SRB in the treatment of AMD. However, most large scale applications of the approach have 

focused on ex-situ treatment in bioreactors or interceptions of contaminated flows such as 

anoxic limestone drains and successive alkalinity producing systems. King et al. (1974) 

describes the recovery of acid strip mine lakes through the natural accumulation of organic 

matter. On this basis they advocated the acceleration of the natural process through additions 

of bulk organic matter such as sawdust, wheat straw, newspaper, manure, and wastewater 

sludge. In-lake neutralization via sulfate reduction is expected to play a key-role in the 

remediation of acidic mining pit lakes (Kleeberg, 1998).


A review by Gibert et al. (2002) found that organic matter type was the prime determinate of 

the efficacy of SRB treatment systems, providing both a suitable redox environment and 

carbon source (Harris & Ragusa, 2000, 2001; Gibert et al., 2002; McCullough et al., 2006). 

The availability of carbon from plant matter is dependent upon the rate of decomposition, 

which can be extremely limited in acidic and anoxic conditions (Harris & Ragusa, 2001). 

Castro and Moore (1997) and Hard et al. (2003) observed that in many sites, the viability of 

this approach required an organic matter source that was effective, economical and locally 

available. This issue is particularly acute in remote mine sites, which prompted Lund et al. 

(2006) to trial Local Council mulch for remediation of a regional Australian pit lake. In many 

remote mining locations, the only bulk carbon sources likely to be available in large 

quantities are sewage from the site or support town, and green waste (a broad range of plant 

material collected through clearing of areas on the mine or in the towns from domestic and 

municipal lawns and gardens and also plant material from clearing of native vegetation as 

part of open cut mining process) (McCullough et al., 2006). Testing similar bulk materials in 

permeable reactive barriers, Waybrant et al. (1998) found they successfully reduced sulfate, 

with sewage sludge the fastest to achieve high levels of sulfate reduction. Testing in 

bioreactors, Harris and Ragusa (2000) also found a mixture of sewage sludge and plant 

material (fresh rye grass) was effective in laboratory experiments at increasing pH (2.3 to 

>3), reducing acidity and divalent metal concentrations of AMD waters through sulfate 

reduction within 30 days. The combined mixture proved more effective in ameliorating pH 

and metal concentrations than either sewage sludge (little response) or fresh plant material 

(nil response) on their own. 

Potential problems with in situ treatment of AMD affected pit lakes include Postgate’s (1984) 

assertion that SRB activities only occur at a pH>5. However, Herhily and Mills (1985) found 

SRB activity in a lake at pH 2.5, and Gyure et al. (1987) in a lake at pH 2.7. McCullough et 

al. (2008) recorded remediation of pH 2.4 pit lake waters by SRB activity in a sewage 

evaporation pond. It appears that less acidic microenvironments in the system can allow SRB 

activity to occur and this activity sustains and increases the size of the microenvironment 

(Küsel & Dorsch, 2000; Küsel et al., 2001). Peine et al. (2000) found that in German pit lakes 

that acidification of the lakes was maintained by the establishment of an acidity driven iron 

cycle. This cycle is dependant on the formation of the mineral Schwertmannite, constant 

38

39 


input of Fe 2+ and no SRB activity below pH of 5.5. In this model, the benefits of addition of 

organic matter to the sediment would eventually be overcome as the cycle became 

established. Other issues that might reduce the effectiveness of this approach to treatment are 

large inputs of acidity from the catchment. 

Very little bioremediation work has been carried out on mining pit lakes in Australia 

(Harries, 1997). As an arid continent, the effects of drought and climate change pose 

particular challenges for mining companies as heavy water users such that in some areas 

water is becoming as valuable as the ore (P. Lilly, AUSIMM, pers. comm.). Therefore there 

is a growing demand for new sources of water to meet a variety of end uses (Doupé & 

Lymbery, 2005; McCullough & Lund, 2006). 

At CCP, water availability for the operation is a priority concern due to competing demands 

for the main water source (Bowen River). Treatment of pit lake water even to only marginal 

quality would reduce pressure on other sources of water (including the Bowen River) for uses 

such as dust suppression (McCullough and Lund, 2006). 

In Chapter 6, we demonstrated in a pilot study in Collinsville that sewage and greenwaste in 

combination had the potential to successfully treat GAE mine lake water. The aim of 

Experiment 2 in Chapter 7, which was conducted in Perth (under more controlled conditions) 

was to demonstrate that the positive effects were reproducible, determine whether reduced 

quantities of organic material would be equally effective and investigate the mechanisms 

responsible for the improvements more closely. Furthermore in Experiment 1 the sewage 

sludge used was from Collinsville Waste Water Treatment Plant. Due to the low production 

from the Collinsville Plant, the sludge had been baked in the sun for several months. In 

Experiment 2 we wanted to test a combination of sewage from Bowen and Collinsville, as 

Bowen sewage is wetter and larger quantities are available.


7.2 Methods 

Experiment 2 consisted of 21 microcosms, with 3 replicates for each of 7 treatments. The 

treatments included an untreated control, and additions (100 g or 200 g) of greenwaste, 

sewage or greenwaste and sewage combined (Table 7.1). However, one replicate of treatment 

S200 failed soon after establishment leaving only two replicates for this treatment. Sewage 

consisted of 1 part of Bowen waste to 2 from Collinsville. 

Table 7.1. Experimental design for organic dosing of pit lake cores. 

Treatment Code Individual 

organic mass 

(g) 

40 

Water:greenwaste:sewage 

ratio 

Control Control 0 1:0:0 

Greenwaste (G100) 100 32:1:0 

(G200) 200 16:1:0 

Greenwaste and Sewage (GS100) 100 32:1:1 

(GS200) 200 16:1:1 

Sewage (S100) 100 32:0:1 

(S200) 200 16:0:1 

Experiment 2 microcosms were made from clean 100 mm diameter and 600 mm long (4.5 L) 

acrylic tubes, sealed at the base with a rubber bung, and at the top with a loose-fitting PVC 

cap which allowed for limited gas exchange between the water and surrounding air. Although 

GAE is thermally stratified for several months each year during which the hypolimnion 

becomes anoxic, for the remainder of the year it is well mixed and oxic (author’s unpublished 

data). The loose covers attempt to simulate reduced mixing and resultant hypoxia caused by a 

thick layer (0.5 to 1 m) of greenwaste added to GAE. Littoral sediment from GAE was added 

to a depth of 140 mm in the microcosm, this sediment had a similar pH in the interstitial 

water to that of the overlying water when collected. A 10 mm diameter fibreglass mesh 

(flywire) tube was placed in each microcosm to permit sampling of bottom waters once 

organic matter was added (Figure 7.1). The microcosms were then filled to within 20 mm of 

the top with GAE water.

41 


The lower 300 mm of the microcosms were placed in an opaque black plastic tub which was 

filled with tap water to evenly distribute ambient temperatures between cores and to reduce 

the possibility of leakage. The tubs were then wrapped to the top of the microcosms in black 

PVC sheeting to occlude light, as PAR levels are extremely low in the hypolimnion of these 

lakes (Author’s unpublished data). 

Liquid sludge was collected from Bowen and primary-treated sewage sludge was collected 

from drying beds of the Collinsville and Bowen Municipal Wastewater Treatment Plants and 

analysed (as per Chapter 6). Homogenised samples of sewage and greenwaste were pH paste- 

tested (water:solid ratio of 10:1). Samples were also analysed for total nitrogen, total 

phosphorus and total organic carbon content and a range of solutes. 

On 10 occasions at increasing intervals, a Hydrolab Datasonde 4a multiparameter meter was 

used to measure temperature, pH, specific conductance (EC), oxidation-reduction potential 

(ORP, platinum reference electrode), and dissolved oxygen (DO, % saturation and mg L -1 

) in 

the water of the microcosm.


Figure 7.1. Microcosm design showing placing of sediment, organic matter and mesh tube for an S200 

treatment. 

A 60 mL water sample was taken 180 days after dosing, filtered through 0.5 µm glassfibre 

filterpaper (Pall Metrigard) and kept at 4 o C prior to analysis for a range of solutes, total 

filtered nitrogen (TFN) and total filtered phosphorus (TFP). 

Microcosm sediment geochemistry was sampled on day 230 by removing unsolidified 

surface material and then homogenizing the remaining upper 20 mm of sediment of each 

microcosm core. Samples were analysed for a range of solutes. 

Sulfate reducing bacteria abundances for each treatment were semi-quantitatively assessed at 

180 days by pooling within a treatment 5 mL aliquots of surface water from each microcosm 

in a “SRB-BART” which measured SRB activity (Droycon Bioconcepts Inc.). SRB-BART 

vials were incubated at 25 o C for 9 days with assessments of their activity made daily. 

42

43 


Sewage, greenwaste, and sediment samples were digested using aqua-regia as per (USEPA, 

1995). The digestate and water samples were analysed by Inductively Coupled Plasma 

Atomic Emission Spectrophotometry (ICP-AES) for Al, As, B, Ca, Cd, Cr, Co, Cu, Fe, Pb, 

Mg, Hg, S, Se, Si, Sn, Zn. Total Kjeldahl N and Total P were measured in sewage and 

greenwaste samples as per APHA (1998). Total Organic C was measured in sewage and 

greenwaste samples on a TOC Analyser 700 (OI Corporation). Water samples were subject to 

a persulfate digestion and analysed for TFN and TFP as per APHA (1998). All analyses were 

undertaken at NATA accredited commercial laboratories. 

The potential for human health risks associated with a field trial of the approach were 

assessed by measuring faecal coliforms and gas production. Faecal Coliforms were measured 

in the water of the microcosms on day 180 by the standard total coliform membrane filter 

procedure for total members of the coliform group (APHA, 1998). Sample blanks and 10 mL 

samples of microcosm surface water were taken with sterile 10 mL syringes, filtered with 

20 mL sterile buffered water onto sterile gridded filters and then placed upon sterile 

rehydrated agar pads staining for total coliforms (Sartorius AG 14053 N, ‘endo media’). 

Coliform colonies were counted following 24 h dark incubation at 35 oC. The concentrations 

of emitted gases (ammonia and hydrogen sulfide) were measured in the capped headspace on 

days 8, 74 and 125. A single sample (400 mL) was collected from each replicate for controls 

(only 3 sampled) and 200g treatments and combined for a total of 1.2 L which was pumped 

through Draeger tubes for both gas types (levels of detection to 0.8 ppm). 

7.2.1 Statistical analyses 

Differences between treatments’ chemistry was determined by ANOVA with pair-wise 

differences determined by Tukey’s Honestly Significant Difference (HSD) post hoc test 

(Tukey, 1953). Multivariate physico-chemical data from day 180 was analysed by ANalysis- 

Of-SIMilarity (ANOSIM) (Clarke, 1993) and Principal Components Analysis (PCA) in the 

PRIMER 6 software package (PRIMER-E Ltd, 2001).


7.3 Results 

On 10/11/2004 Garrick East pit lake had a very low pH (2.41 - 2.33), high EC (9.04 - 

9.26 mS cm -1 ), high ORP (594-623 mV), and thermally stratified (temperature changes over 6 

m from 27.1 to 21.2 o C) and anoxic in the hypolimnion (DO surface - 7.37 mg L -1 , bottom - 

0.03 mg L -1 and high total iron concentrations (Table 7.2) indicative of a strongly iron 

buffered system (Totsche et al., 2003), with heavy metals at environmentally toxic 

concentrations (ANZECC/ARMCANZ, 2000b). Nutrient levels were also very low in this pit 

lake; especially phosphorus, as is typical of many AMD waters (Lessmann et al., 2000; Borg 

& Holm, 2001). This water was used to create the microcosms and is a reference condition 

for the controls (Table 7.2). By Day 180, the effects of the microcosm on water quality in the 

absence of treatment can be assessed through comparison of the original GAE water and the 

control. Sulfate levels increased substantially and nitrogen appeared to decline substantially 

(although the change from TN to TFN could account for the change). All other parameters 

changed little in concentration (Tables 7.2). pH ranged between 2.15 and 2.45 (mean±s.e.; 

2.34±0.009), EC ranged from 7.24 to 10.42 mS cm -1 (9.48±0.06 mS cm -1 ), DO ranged 

between 3.0 mg L -1 (37% saturation) and 7.4 mg L -1 (93%) (5.5±0.1 mg L -1 ) and ORP ranged 

between 534 and 657 mV (601±3 mV). The water temperature for all the microcosms ranged 

between 20.9 and 27.5 o C (25.0±0.1 o C). 

44

45 


Table 7.2. Chemistry of GAE pit lake water (April 2006) at collection and control microcosms at Day 180. 

Parameter GAE water 

(mg L-1) 

Day 180 

Controls 

(mg L-1) 

Sulfate 2,610 9707±942 

Total N 0.51 0.5±0.2 (TFN) 

Total P


Table 7.3. Chemistry of the sewage and greenwaste used in the experiment. 

Parameter Collinsville 

sewage 

(mg Kg -1 ) 

46 

Bowen 

sewage 

(mg Kg -1 ) 

Greenwaste 

(mg Kg -1 ) 

Total sulfur 10,000 14,000 1,700 

Total N 31 П 60 П 20 

Total P 12 19 2.3 

Total organic C 29% 34% 39% 

pH 6.1 6.5 5.6 

Aluminium 17,000 7,600 1,000 

Arsenic 7 7 1 

Cadmium 3.2 2.1

47 


Table 7.4. Maximum increase in GAE water solute concentrations from each treatment in the microcosms 

with reference to initial GAE water quality (– = No data; *Total nitrogen as Kjeldahl nitrogen). 

Maximum potential addition to GAE water concentration (mg L -1 ) 

GAE Greenwaste Sewage Greenwaste & Sewage 

(mg L -1 ) 100g 200g 100g 200g 100g 200g 

Total sulfur – 24 47 266 531 145 289 

Sulfate 2,610 – – – – – – 

Total N 510 280 560 1,060 2,120 670 1,340 

Total P


improvements in pH, with greenwaste and sewage combined treatments providing the 

greatest improvement, followed by sewage and then greenwaste alone. Adding more organic 

material, with the exception of greenwaste only, improved the pH reached at day 180. For 

example, GS100 produced a similar pH and water quality to S200. Improvements in pH were 

associated with reductions in S, sulfate, ORP, DO, Al, Cd, Co, Cr, Ni, Zn, Cu, Pb, and Fe 

concentrations, which are consistent with sulfate reduction processes. The heterogeneous 

nature of the greenwaste appears reflected in the variability between replicates which was 

larger than for other treatments especially at the 200g level. ; Interestingly, S100 was 

associated with higher Se, TFN, TFP, Mg and Ca concentrations compared to all other 

treatments. The addition of more organic matter (200g) to the treatments, resulted in lower 

EC, despite potentially adding more solutes (see Table 7.4). 

Figure 7.2. PCA of day 180 microcosms physico-chemical data. The amount of variability explained by each 

axis is shown as a percentage. Lines indicate eigenvectors, the direction of the line indicates the 

relationship to a PC axis and the length of the line indicates its relative importance (longer lines are 

more important). 

48

49 


Day 180, total nutrients increased from below detection (


Table 7.6. Mean (± standard error) of selected major cations in water from microcosms collected on Day 180 

(no standard error is shown for S200 treatment as there were only two replicates available) 

Microcosms Al Ca Fe K Mg Mn Na Zn 

mg L -1 mg L -1 mg L -1 mg L -1 mg L -1 mg L -1 mg L -1 mg L -1 

Control 142.5±8.5 398±20 468±156 0.8±0.5 550±19 47.3±1.5 408±13 13±0.7 

G100 56.5±34.1 470±0 560±50 147±30 577±17 46.3±0.9 533±15 4±3.5 

G200 65.8±32.3 463±20 284±234 179±59 563±18 42.7±2.0 470±21 6.5±3.4 

G100 0.02±0.01 400±11 24±23 47±0.7 583±26 30.7±3.3 577±32 0.08±0.04 

G200 0.02 395 0.2 82 565 3.2 710 0.05 

GS100 0.2±0.1 367±35 112±109 167±12 526±26 5.5±0.5 600±6 0.07±0.006 

GS200 0.01±0.01 52±4 0.5±0.1 333±20 443±23 0.7±0.1 773±71 0.18±0.06 

Table 7.7. Mean (± standard error) of selected minor cations in water from microcosms collected on Day 180 

(no standard error is shown for 200 g Sewage treatment as there were only two replicates 

available. *contains replicates that had concentrations below detection levels – As & Pb

51 


3.0 mg kg -1 in control to 2.1 mg kg -1 in GS200. Conversely, B concentrations were raised in 

greenwaste and sewage-dosed treatment sediments; from 5.9 mg kg -1 in control to a mean 

maximum of 8.5 mg kg -1 in GS200. Mean Ca concentrations were raised from 1.7 g kg -1 in 

control in all treatments other than GS100 where it was reduced to 1.3 g kg -1 . Mean Co 

concentration was greatly increased in G treatments from 2.5 mg kg -1 in control to a mean of 

12 mg kg -1 in G100. Mean Cr also increased in all treatments from 2.0 mg kg -1 to means of 

2.7 and 2.8 mg kg -1 in GS100 and S100 respectively. Mean Cu concentrations increased in 

proportion to the amount of sewage in the microcosm to a mean of 21 mg kg -1 in S200. Mean 

Fe sediment concentrations decreased in all dosed treatments from 9.1 g kg -1 in control to a 

mean minimum of 6.9 g kg -1 in GS200; greatest reductions occurred in treatments containing 

sewage. Mean K, Na , Pb and S concentrations changed little across treatments. Mean Mg 

sediment concentrations increased in all treatments from 602 mg kg -1 in the control to a 

maximum of 1.1 g kg -1 in S100, except for GS200 where it decreased to a minimum of 

563 mg kg -1 . Mean manganese concentrations increased from 65 mg kg -1 in the control to 

100 mg kg -1 in both G100 and S100, but reduced to only around 42 mg kg -1 in GS treatments. 

Sediment Ni concentrations decreased from 4.6 mg kg -1 in all sewage containing treatments 

except for S100, but increased in both G treatments. Ni concentrations were lowest in MS200 

at 2.9 mg kg -1 , and highest in G100 at 20.3 mg kg -1 . Sediment concentrations of Zn also 

increased in all treatment, from a mean control concentration of 37 mg kg -1 to a maximum of 

79 mg kg -1 in S100.


Figure 7.3. Day 230 concentrations of a). major, and b). minor, sediment analytes for each organic matter type 

and dosing rate. Error bars indicate single standard errors of the mean. 

52

7.3.3 Changes in water quality during the experiment 

53 


Temperature was very similar within treatments, ranging from 23.9–27.4 o C and averaging 

25.0 o C (Figure 7.4). The coefficient of variation for temperature over time and between all 

treatment replicates was only 5%. The pH of the control treatment level changed little (0.05 

pH units) over the course of the experiment. pH was not seen to increase as much in either G 

treatments, with a maximum of pH 4.5 reached by final day 228. However, sewage, and 

green waste and sewage treatments both showed a rapid increase in pH to circum-neutral pH 

by day 282, except for S100 which began to plateau at pH 6. There also appeared to be 

greater variation in the pH response within the G treatments than for either S or GS 

treatments. 

Specific conductance appeared to increase in all treatments following addition of organic 

material with the greatest increase in the green waste treatment (Figure 7.4). Although there 

was a slight increase in control specific conductance over the experiment from 9.4– 

9.9 mS cm -1 

, specific conductance appeared to only decline in both greenwaste treatments and 

S100 at around the same as alkalinity increased (i.e., around day 37). Although treatments 

containing sewage were very variable between replicates, both treatments containing sewage 

appeared to decline in solute concentrations greater than did green waste alone.


54

55 


Figure 7.4. Change in a). pH, b). EC, and c). ORP of microcosm treatment levels over time for different 

organic dosing treatments. Error bars indicate single standard errors of the mean. 

Although there was generally little different between upper and lower pH of microcosms, 

S200 had higher pH near the sediment than near the water surface in the first 30 days 

following dosing. Conversely GS100 had higher pH near the water surface than near the 

sediment in the first 90 days following dosing. However, after 60–90 days, the difference in 

pH between upper and lower of all treatment microcosms was less than 1 pH unit (Figure 

7.5). 

Figure 7.5. Treatment pH difference (top pH - bottom pH) of microcosms over time for different organic dosing 

treatments. Error bars indicate single standard errors of the mean. 

7.3.4 Bacterial abundances 

Faecal coliforms were not observed in any treatments. However, total coliform counts were 

too abundant to be enumerated in all treatments amended with greenwaste. Sulfate reducing 

bacteria activity was only found in the organic-supplemented treatments containing sewage 

(GS100, GS200, S100 and S200). A black precipitate also formed immediately upon mixing 

in all of these treatments, except S100, indicating excess dissolved H 2S.


7.3.5 Gases evolved 

Hydrogen sulfide was also evolved as the black precipitate extended upwards through the 

microcosm waters above the green waste and sewage solids. The sulfide smell and the 

concomitant presence of a black precipitate indicated that sulfate reduction was likely 

occurring in all treatments containing greenwaste. There were no detectable traces of 

ammonia or hydrogen sulfide on day 8. However, by day 180 there were 74 ppm NH 3 and 

12 ppm H 2 S in GS200, and 17 ppm H 2 S in S200. By day 125 there remained 12 ppm NH 3 in 

GS200, and 7 ppm of NH 3 and H 2 S in S200. As the microcosms cores were loosely sealed at 

the top, these concentrations are indicative only of the nature and relative rates of 

biochemical processes taking place in the cores. 


The decrease in sulfate concentrations in all treatment microcosms provides support for 

biological sulfate reduction producing alkalinity and removing sulfate and cations as 

monosulfides. The success of sulfate reduction in these low pH waters appears to be contrary 

to Postgate’s (1984) (amongst other authors) assertion that sulfate reducing bacteria (SRB) 

activities only occur at pH >5. Nevertheless, other published studies have reported SRB 

activity at pH values below this threshold. For example, SRB activity in an acid lake has been 

found by Herhily and Mills (1985) at pH 2.5, and at pH 2.7 by Gyure et al. (1987). It appears 

that mechanisms of adaptation and/or less-acid microenvironments, enabled sulfate reducing 

bacteria to begin reducing sulfate, promoting alkalinity as part of this biological process 

(Küsel & Dorsch, 2000; Küsel et al., 2001). 

Variability within control sediment metal concentrations was likely to be due to natural 

sediment heterogeneity. However, contrary to that expected by sulfate reduction, sediment 

total sulfur concentrations were not necessarily increased where water column concentrations 

of sulfate were greatly reduced. However, unlike oxic lakes, pyrite has also been found to 

form in the anoxic water column of natural water bodies (i.e., syngenetic pyrite) due to 

56

57 


limited availability of reactive iron in benthic sediments, (Suits & Wilkin, 1998). 

Consequently, although the intention of removing organic sludge from the sediment surface 

was to prevent confounding by remaining organic matter, it is likely that amorphous metal 

monosulfides were also removed by this activity, confounding sediment results. 

All treatment faecal coliform abundances were below detection level of 10 MPN 100mL -1 

, 

which met ANZECC/ARMCANZ (2000a) contact limits of 150 faecal coliforms 100 mL -1 

. 

However, the extremely high total coliform count (greater than could be enumerated) may 

mean that further organic decomposition than 180 days is required before contact may be 

possible in pit lakes remediated with greenwaste. Nevertheless, the presence of equally high 

abundances of total coliforms in greenwaste-only samples indicates that they are likely to be 

simply non-pathenogenic decomposers that occur in benthic sediments of many natural water 

bodies. 


(Berrow & Webber, 1971). Both the Collinsville and Bowen sewage displayed high 

concentrations for many heavy metals. However, these heavy metals are also likely to be 

trapped in the sediments as metal sulfides due to SRB activities. Careful management of the 

remediated pit lake would then be required to prevent oxidation of these sediments and 

subsequent remobilised (Simpson et al., 1998). 

Further reduction in biological availability of metals and metalloids is also likely to occur 

through formation of complexes with organic chelators present as components of the 

refractory green waste, such as organic acids (Tipping & Hurley, 1992). Nevertheless, the 

contribution that organic materials may make to the heavy metal burden of a pit lake needs to 

be considered in the choice of organic materials for remediation. This study found sewage 

treatment performed similarly to that of greenwaste and sewage, whilst others have found 

vegetation has performed equally (Waybrant et al., 1998; McCullough et al., 2006). In this 

respect, remediation strategies may be best placed by choosing greenwaste as the bulk 

contribution to electron donors over that of sewage. As discussed, greenwaste also has an 

additional advantage of providing organic substances such as humic and fulvic acids, with 

which heavy metals may directly complex to. Ligand formation between heavy metals and


refractory organics will further remove these toxic components from biological availability, 

albeit at a likely reduced capacity to that of sulfate reduction processes (Brown Jr., 2001). 

Nevertheless, a clear concern of the use of sewage is the introduction of human pathogens. 

The coastal town of Townsville, only 170 km to the north-north-west of Collinsville, receives 

a very high mean annual ultraviolet (UV) exposure of 4 300 J m -2 , and a maximum of around 

60 J m -2 in December (personal communication John Javorniczky, Australian Radiation 

Protection and Nuclear Safety Agency), indicative of the high UV levels experienced in 

Collinsville pit lake surface waters. Consequently, deactivation of coliforms in the warm 

saline waters of the pit lakes under this high level of UV exposure is likely to be very rapid 

(Smith et al., 1994; Sinton et al., 1999), occurring within days. Furthermore, even if bacteria 

remain unexposed in sediments and deeper waters, our mesocosm results (conducted under 

nil UV exposure) indicate that faecal bacteria are not able to survive for extended periods 

away from host organisms and in acidic waters. 

Although lake release of ammonia gas is unlikely to be significant, release of toxic hydrogen 

sulfide from the water body may occur when dissolved concentrations of iron are 

significantly decreased. Nevertheless, “run away” sulfate reduction may be feasibly 

controlled by introducing more low pH iron-laden AMD water to react with excess hydrogen 

sulfide and lower water pH to that sub-optimal for sulfate reduction bacteria. These findings 

were incorporated into a formal risk assessment for the field-scale experiment (Chapter 

8.2.1). 

Consequently, this study has demonstrated remediation of AMD water pH and high metal 

concentrations within months of dosing with green waste and sewage. This work builds on 

other work demonstrating that readily available and cheap organic matter may be used to 

develop alkalinity though microbially-mediated in situ sulfate reduction. For example, 

Waybrant et al. (1998) tested eight bulk organic matter types, with sewage sludge fastest to 

achieve high levels of sulfate reduction. Again using locally available bulk materials, Harris 

and Ragusa (2000) also found a mixture of sewage sludge and plant material (fresh rye grass) 

effective in remediating acidity and high metal concentrations of acid mine waters through 

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59 


sulfate reduction. Different to this current study, their combined mixture proved more 

effective in ameliorating pH and metal concentrations than either sewage sludge (little 

response) or plant material (nil response). Conversely, an observation at the Collinsville Coal 

Project Mine in North Queensland, Australia (Fallon, 1994) suggested that reduced acidity, 

sulfate and metal concentrations in a pit lake was due to bacterially-mediated sulfate 

reduction with the sewage effluent discharged into this pit functioning as a carbon source. 

In conclusion, mine water research and management is a very new and rapidly developing 

field (Wolkersdorfer, 2004) and corrective measures for acidic mining lakes also greatly 

differ from those in use for eutrophication control (Klapper, 2003). Best practice AMD pit 

lake treatment is likely to be required to be made on a case-by-case basis where potential pit 

lake end uses and availability of organic substrates have been identified and a remediation 

strategy has consequently only then been constructed. Consequently researchers, consultants 

and regulatory agencies need to maintain an open-mind to treatment solution options for 

these unique environmental issues. Part of the challenge for both mining companies and 

regulatory waterbody managers with this assessment will be the ability to think laterally as to 

what the current values of the pit lake are, what treatments may be feasible, and what type of 

pit lake (water quality and quantity) is desirable for either social, environmental or other 

enduses (Doupé & Lymbery, 2005; McCullough & Lund, 2006).

61 


8 Field-scale remediation of a tropical acid mine pit 

lake with greenwaste and sewage sludge 


The last half century has seen development of technologies resulting in large-scale open-cut 

mining in areas which would previously have been worked underground, if mined at all. This 

activity has left a legacy of many thousands of mine pit lakes worldwide (Klapper & Geller, 

2002). As these technologies continue to develop, the result is likely to be ever larger and 

deeper mining pits. Most mine pits are abandoned as ‘open voids’, with the burden and 

tailings left outside of the pit as this is the cheapest and most practicable option for mining 

companies (Castro & Moore, 1997). This type of mine closure allows easier access to re-mine 

tailings and lower grade ores if technology or economic change renders this feasible. 

However, open voids may also be the only practical option for very large pits which are not 

directly linked to significant regional water resources and are too large to even partially 

backfill; either at completion or during active working. 

Pit lakes form in these voids when dewatering ceases in a mine pit that extends below the 

watertable. As groundwater, rainfall and surface runoff slowly fills the pit, this water may 

react with oxidised seams and pit walls resulting in dissolution and oxidation of exposed 

minerals (Castro & Moore, 1997). Pit lakes may be terminal groundwater sinks when the 

climate is net-evaporative, with the pit lake creating a localised depression in the surrounding 

water table. Under these net-evaporative conditions, soluble ions may increase in 

concentrations by evapo-concentration (Commander et al., 1994). Alternatively, under a net- 

precipitation regime, pit lakes may contribute as sources to local hydrology with either 

surface or ground water outflow or may function as flow-through “groundwater windows” 

(Johnson & Wright, 2003). 

In these two latter cases of net precipitation, the open void strategy may have long-term 

benefits or liabilities to associated communities and the natural environment (McCullough et 

al., in review). Whilst mining companies and communities have often failed to appreciate the


benefits and opportunities of such a large waterbody in post-mining landscapes, various 

opportunities frequently exist for pit lakes and their water (McCullough & Lund, 2006). 

These include water for commercial operations such as tourism, aquaculture, tree farming and 

horticulture; as well as recreational opportunities such as swimming, fishing, water skiing, 

and wildlife observation. Nevertheless, pit lakes of low water quality may threaten the health 

and well-being of both local communities and the natural environment (Doupé & Lymbery, 

2005). 

Acid Mine Drainage is the most common problem affecting pit lake water quality globally 

(Charles, 1998; Klapper et al., 1998). Receiving environments for AMD typically have 

reduced environmental and social values, and the resultant water is less valuable as a resource 

to the mining company (McCullough & Lund, 2006). Acid Mine Drainage is arguably the 

greatest environmental problem facing water management in the international mining 

industry today (Gray, 1997; Harries, 1998a). The large quantities of water in pit lakes 

potentially represents a potentially valuable resource to mining companies, the environment 

and community; if appropriate water quality can be achieved (Doupé & Lymbery, 2005; 

McCullough & Lund, 2006). For example, the Bowen region in North Queensland of 

Australia is currently facing a long-term drought and the large volumes of water, even of low 

quality, are of great benefit (Côte et al., 2006). Mining lease pit lakes therefore represent 

significant short-term resources to adjacent operations, and furthermore reduce pressure on 

regional natural water resources over a longer period (McCullough & Lund, 2006). 

Microbial sulfate reduction can be an efficient and effective remediation for the treatment of 

AMD contaminated waters. Sulfate reducing bacteria (SRB) can reverse the acidification 

process by converting sulfates to sulfides in low redox environments when supplied with 

labile carbon sources (see King et al, 1974). Tuttle et al. (1969) first suggested the use of 

SRB in the treatment of AMD. Such in-situ neutralisation by sulfate reduction is expected to 

play a key-role in the remediation of acidic mining pit lakes (Kleeberg, 1998). However, 

primarily due to carbon and P limitation, acid mine lake tropic status is ultra-oligotrophic to 

oligotrophic (extremely low in nutrients) (Lessmann et al., 2000; Nixdorf et al., 2005) and 

acid autogenic carbon accumulation in is typically very low (Kleeberg & Grüneberg, 2005). 

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63 


In Experiments 1 and 2, we tested sewage and greenwaste in microcosms for the treatment of 

AMD waters from a Collinsville Coal Project pit lake (GAE). Results were extremely 

positive with rapid remediation from pH 2.2 to >5.5 over 145 days. However, the application 

of this science now required on-site experiments at lake-scale to demonstrate this new 

technology of remediation with bulk low-grade organic sources. This chapter reports on a 

field-scale experiment that aimed to test the suitability of greenwaste and sewage sludge as an 

organic matter source for in situ AMD pit lake remediation. 

8.2 Methods 

The proposed experiment underwent an extensive and comprehensive risk assessment and 

occupational health and safety assessment prior to being approved. These assessments were 

led by the author (JM) and is detailed in Chapter 8.2.1. 

Previous microcosm experimentation with an AMD:sewage:greenwaste ratio of 16:1:1 had 

been successful in remediating GAE water to pH 5.5 in under 5 months (see Chapter 7). Due 

to the large size of the GAE pit lake, it was divided by earthworks in July 2006 into a smaller 

eastern lake (GAEE), to be treated, and a larger western (GAEW) control (untreated) lake. 

This was to ensure that the organic loading required could be achieved over six-months 

(Figure 8.1). In addition to GAEE, three representative pit lakes on the lease were used as 

further controls (Ramp 8, Ramp 5 and Ramp 5a). Operational reasons prevented continued 

monitoring of control Ramp 5 after 3 months of data collection . The small dataset that had 

been accrued during this time was therefore not used in further analysis. Ramp 8 also became 

inaccessible before project completion, however 20 months of data had been collected and 

was this was used in analysis. Although Ramp 5a is shallower, Ramp 8 was comparable to 

GAEE (Table 8.1).


Figure 8.1. GAE after being divided into two new lakes; GAEE and GAEW. 

Table 8.1 Morphology of project control and treatment lakes. 

Pit lake Maximum depth (m) Surface area (ha) Volume (ML) 

Ramp 5 4 4.6 404 

Ramp 5a 5 3.4 117 

Ramp 8 12 1.2 95 

GAE 13.8 5.9 470 

GAEE 13 4.4 336 

GAEW 12 1.1 71 

64

8.2.1 Risk Assessment 

65 


The location and nature of the field-scale trial necessitated extensive consultation and 

assessment from a number of stakeholders prior to the commencement of the trial. A list of 

the key stakeholders and their input into the project is shown below: 

Bowen Shire Council: supported the project in providing the organic materials 

(sewage sludge and green waste) for the project; 

Queensland Environmental Protection Agency: provided state government 

environmental approval for the research project; 

Mine workforce: a presentation was made to representatives of the local CFMEU 

lodge, and a cross-section of workforce was engaged in the formal risk assessment 

process; 

Local community groups: a presentation was made to the Retired Miners’ 

Association, and newsletters mailed to the local communities of Collinsville and 

Scottville which provided a description of the proposed project and subsequent 

updates seeking feedback from the community; and, 

Industry organisations: presentations were made to local, regional and international 

industry organisations and feedback received on aspects of the project. 

The outcome of these consultation sessions involved the development of a specific project 

work procedure which detailed the requirements at various stages of the project including 

communication, occupational health and safety requirements. These requirements are 

summarised below: 

Communication requirements: 

o external communication involved project updates to key stakeholders, and also 

specific communication at particular stages of the project (i.e., Bowen Shire 

Council and the waste transport contractors were regularly contacted during 

the addition of waste stage);


o internal communication involved coordination of project activities within 

general mining operations (i.e., communication with Open Cut Examiner 

when work was being conducted in Garrick East pit area, restricting the 

project site to project personnel as part of general mine safety procedures). 

Occupational health and safety: 

o Due to the nature of the material (especially the sewage sludge) that was 

involved in the project, all personnel involved in the transport of sludge waste 

and contact of waters which contained this material as part of the monitoring 

program, underwent appropriate immunisation (i.e., hepatitis injections) and 

were required to follow a hygiene protocol; 

o As the laboratory core experiments had detected some gas emanating from the 

8.2.2 Organic dosing 

cores, gas monitoring was conducted by both mine and project specific 

personnel at regular intervals. This monitoring was supported by a trigger 

action-response plan (TARP) as part of the work procedure. Similarly, project 

specific and other general mine emergency procedures were linked to provide 

an effective framework for dealing with an emergency. 

Over six months from August 2006 to February 2007, primary-treated sewage sludge and 

liquid was collected from the Bowen Municipal Wastewater Treatment Plant and green waste 

was collected from the Collinsville Municipal Landfill. The quantity of organic added was 

obtained from estimates of volumes, and number and volume of cartage trips for all organic 

matter types and from sewage and greenwaste density data from Experiment 2 and for Bowen 

liquid sewage from a calculated specific density of 1.02 kg L -1 . From July–December 2006 

approximately ca.440 t of liquid sewage was added to GAEW per month. Sludge drying rates 

were reduced during the wet season, therefore this dosing rate was reduced from January– 

February 2007 to ca.290 t per month. Dried Collinsville sewage sludge was added as 20 t in 

November 2006 and 40 t in January 2007. Greenwaste was added as ca.730 t and ca.250 t in 

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67 


August and October 2006 respectively. This 13:1:1 ratio of sewage:greenwaste:water 

represented around 80% of the most effective laboratory-derived dose of 16:1:1 (see Chapter 

7), with greenwaste the limiting available organic material. Greenwaste was largely 

representative of Australian garden waste and consisted of lawn clippings, tree branches and 

palm fronds, and other leafy and woody material. Sewage was deposited from 4 points 

around GAEW with green waste further spread by a long-arm excavator as it was deposited 

(Figure 8.2). 

Figure 8.2. Green waste being spread across the surface of treatment lake GAEW. Control lake GAEE is in 

the background (Photo: Joel May, Xstrata Pty Ltd). 

A sample of the sewage sludge was digested after APHA (1998) and analysed for total 

nitrogen, total phosphorus and total organic carbon content, as well as the cations Al, Ba, Be, 

Ca, Cd, Cr, Co, Cu, Fe, Mn, Na, Ni, Sb, Sn and V by Inductively Coupled Plasma Atomic 

Emission Spectrophotometry (ICP-AES). Representative chemical data for green waste was 

taken from McCullough et al. (2006).


8.2.3 Sampling 

8.2.4 Groundwater interactions 

Ground water monitoring bores were located immediately to the west (GAR01, 14.5 m deep), 

east (GAR02, 15 m deep) and south (GAR03, 23 m deep) of GAE pit lake (Figure 8.3). 

sampled using a bailer. Monitoring bore water levels was sampled from December 2003 to 

October 2007. GAE and GAEE/GAEW water levels were sampled from July 2004 to October 

2007. From July 2004 till October 2007 monitoring bores were sampled in situ monthly for 

water level, pH, EC and temperature with a Hydrolab Minisonde meter. From November 

2005 to July 2007 groundwater solute samples were taken at quarterly intervals using 

standard sampling protocols and monthly intervals in the second-half of 2006. A water 

sample was kept chilled prior to analysis in a commercial NATA registered analytical 

laboratory for dissolved metals by ICP-MS (Al, As, Be, Bi, Cd, Cr, Co, Cu, Pb, Li, Mn, Mo, 

Ni, Sb, Se, Ag, U, V, Zn) and hydrocarbon and total nutrients. 

Figure 8.3. Location of GAE pit lake monitoring bores. 

68

8.2.5 Pit lakes 

69 


A vertical profile of the water column was taken in the centre of each lake at the end of every 

month between April 2005 and July 2007. Temperature, pH, specific conductance (EC), 

oxidation-reduction potential (ORP, platinum reference electrode), turbidity (NTU), and 

dissolved oxygen (DO, % saturation and mg L -1 

) were measured at 1 m intervals using a 

Hydrolab Quanta multi-parameter meter. Chlorophyll a was measured with a Turner Designs 

Cyclops fluorometer attached to a volt meter calibrated against known chlorophyll 

concentrations in pit lake water. Known chlorophyll concentrations were determined by 

filtering five different concentrations of GAEW pit lake surface waters (0.5 µm Pal 

Metrigard). The filter paper was frozen, and following extraction with Dimethylformamide 

(DMF) (Speziale et al., 1984) chlorophyll a was measured with a Schimadzu UV-1201 

spectrophotometer as per APHA (1998) to derive a calibration curve (r 2 = 0.997). 

A water sample was collected for analysis of nutrients and metals from the lake surface as 

well as another from ca.0.30 m above the benthos using a Kemmerer bottle. Upon collection, 

each water sample was split into one 250 mL aliquot of unfiltered and one aliquot filtered 

through glassfibre filterpaper (0.5 µm Pal Metrigard). Aliquots were stored in acid washed 

high-density polyethylene bottles and stored at 4 o C prior to analysis. Part of the filtered 

aliquot was analysed for SO4 2- by ion chromatograph (Dionex ICS-1000). Another portion of 

the filtered sample was analysed for dissolved organic carbon as gilvin 440 by absorbance on a 

spectrophotometer at 440 nm. The remaining filtered sample was then acidified with reagent 

grade HCl and selected metals/metalloids analysed by Inductively Coupled Plasma Atomic 

Emission Spectrophotometry (ICP-AES; Al, Ba, Be, Ca, Cd, Cr, Co, Cu, Fe, Mn, Na, Ni, Sb, 

Sn and V). Unfiltered samples were digested using a persulfate digestion and then analysed 

by discrete analyser for total P and total N. 

8.3 Results 

8.3.1 Climate 

Australian Bureau of Meteorology data (Australian Bureau of Statistics, 2007) indicate that 

over the sampling period, rainfall was characteristically unpredictable with low rainfall in


2005 until the Wet season rains in January 2006. Late Wet season rains then followed in 

April, followed by low rainfall again until very heavy rainfall episodes in January and 

February 2007. June 2007 was uncharacteristically wet, with no rainfall in July (Figure 8.4). 

Warmer over the late Dry and Wet season months of October through to April, mean monthly 

temperatures were more predictable, albeit slightly warmer on average in 2005–2006 

(27.8 o C) than in 2006–2007 (26.1 o C). Nevertheless, rainy days in June 2007 led to 

particularly cold days, and clear skies in July to particularly cold nights. 

8.3.2 Dosing with organic material 

Figure 8.4. Collinsville climate for April 2005 to July 2007. 

Organic material was dosed in treatment lake GAEW from July 2006 to February 2007. Total 

chemical loadings were calculated from tonnage of loading and concentrations of chemicals 

in organic materials samples (Table 8.2). Unsurprisingly, the bulk of organic loading was 

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71 


with organic carbon (402 t). Total Kjeldahl nitrogen and calcium were the next greatest loads 

at 29 and 28 t respectively. 

Table 8.2. Chemistry and total loading of organic materials dosed. *Data from Experiment 2. – = no data 

available. 

Parameter Bowen sewage liquid Bowen sewage sludge* Green waste* Total loading 

Loading (t) 3,193 61 979 4,233 

Analyte concentrations (mg/Kg) Total analyte load (kg) 

TN 1,800 60,000 20,000 29,000 

TP 440 19,000 2,300 4,800 

TOC 340,000 390,000 402,000 

Al 83 7,600 1,000 1,700 

As


8.3.3 Physico-chemical parameters 

Ramp 5a temperature was around 20 oC and well-mixed during Dry seasons from June to 

September (Figure 8.5a). Warming began at the end of the Dry season from October with 

slight thermal stratification occurring in both sample years around December and lasting to 

June. However, curiously, water temperatures were warmest in the centre of the water 

column, from then until June, when mixing was complete again. The water temperature of 

Ramp 8 (Figure 8.5b) became strongly stratified at around 3 m in September, at least three 

months earlier than Ramp 5a. Stratification was also much stronger, with a sharp temperature 

transition in the first 0.5 m, with the hypolimnion proper beginning at around 6 m. Although 

this lake only barely stratified prior to splitting in mid 2006, GAEE mixed completely from 

May to October each year (Figure 8.5c). Nonetheless, following splitting a strong 

thermocline developed with the epilimnion in the upper 2 m from January to May 2007. The 

western side of GAE (GAEW) mixed and warmed at the same time as the eastern side, 

however a strong thermocline at 2 m developed in this side from September to March 2006 

(Figure 8.5d). Following splitting, this thermocline deepened to 4 m and was maintained until 

June 2007. 

Oxidation-Reduction Potential was very high (ca.600 mV) in all control lakes (Figure 8.6). 

Ramp 8 showed a regular bi-annual decrease in ORP of hypolimnion waters, decreasing in 

both mid-Dry season and mid-Wet season (Figure 8.6b). Sometimes these low ORP waters 

extended right through to the epilimnion. However, following the heavy cyclonic rains of 

December–January 2006, Ramp 5a and GAEE showed large well-mixed decreases to 

450 mV in ORP (Figure 8.6). GAEW also showed a moderate decrease in epilimnion ORP, 

however this pit lake was not well mixed. As early as October 2006, ORP in GAEW was 

much lower in hypolimnion waters, with ORP increasing in value as it neared the lake surface 

and with water column ORP decreasing after dosing to a minimum of 65 mV near the lake 

benthos. 

pH was stable around 2.2 in control lakes except for Ramp 8 where it increased to pH 2.7 

during 2005 and 2006 Dry seasons (Figure 8.7). GAEE pH was as low as control lakes prior 

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73 


to dosing, however, following dosing in August, GAEW hypolimnion pH began to increase 

from December 2006 to a maximum of 4.1 in July 2007. Unlike the second mesocosm 

experiment (Chapter 7) pH was not lower at the water’s surface. 

Although specific conductance (EC) was consistently high across all lakes at around 

15 mS cm -1, specific conductance painted a complex picture over time (Figure 8.8). There 

appeared to be a general trend of increasing EC over time, however this was interrupted, 

across all lakes, by a sudden decrease in EC in February 2006. This interruption coincided 

with heavy January 2006 rainfalls after a prolonged Dry period to at least April 2005 (Figure 

8.8). Further heavy rains occurred in January 2007. The epilimnion of Ramp 5a then formed a 

lower EC halocline at around 9 mS cm -1 through to July 2007. Although EC dropped to 

around 10 mS cm -1 in Ramp 8 it did not form a halocline. Both GAEE and GAEW also 

formed haloclines of around 8 mS cm -1. These two haloclines appeared to mix through and 

disappear by May 2007, although a lower EC epilimnion was still maintained in GAEW. 

Although generally below detection limits, pit lake chlorophyll a also showed high variability 

over time (Figure 8.9). Except for Ramp 8, all lakes showed a 2–3 months period of high 

chlorophyll concentrations up to at the start of monitoring in May 2005. Ramp 8 also 

demonstrated two chlorophyll blooms of 6–8 months each over the Wet seasons of 2005– 

2006 and 2006–2007. Unlike its counterpart GAEE, GAEW showed a whole-water column 

increase in chlorophyll a beginning September 2006, two months after dosing. These 

chlorophyll concentrations gradually reduced in the downwards extent to the monitoring 

completion in July 2007.


(a). 

(b). 

(c). 

(d). 

Depth (m) 

Depth (m) 

Depth (m) 

Depth (m) 

0 

1 

2 

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34 

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32 

34 

5/2005 7/2005 9/2005 11/2005 1/2006 3/2006 5/2006 7/2006 9/2006 11/2006 1/2007 3/2007 5/2007 7/2007 

Figure 8.5. Temperature profiles of pit lakes (a) Ramp 5a, (b) Ramp 8, (c) GAEE, (d) GAEW from May 2005 to 

August 2007. 

74 

Date 

18 

20 

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34 

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(a). 

(b). 

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(d). 

Depth (m) 

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Figure 8.6. ORP profiles of pit lakes (a) Ramp 5a, (b) Ramp 8, (c) GAEE, (d) GAEW from May 2005 to August 

2007. 

100 

200 

300 

400 

500 

600 

700


(a). 

(b). 

(c). 

(d). 

Depth (m) 

Depth (m) 

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12 

5/2005 7/2005 9/2005 11/2005 1/2006 3/2006 5/2006 7/2006 9/2006 11/2006 1/2007 3/2007 5/2007 7/2007 

Figure 8.7. pH profiles of pit lakes (a) Ramp 5a, (b) Ramp 8, (c) GAEE, (d) GAEW from May 2005 to August 

2007. 

76 

Date 

2.0 

2.5 

3.0 

3.5 

4.0 

2.0 

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3.0 

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4.0

(a). 

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5/2005 7/2005 9/2005 11/2005 1/2006 3/2006 5/2006 7/2006 9/2006 11/2006 1/2007 3/2007 5/2007 7/2007 

77 

Date 


Figure 8.8. EC profiles of pit lakes (a) Ramp 5a, (b) Ramp 8, (c) GAEE, (d) GAEW from May 2005 to August 

2007. 

6 

8 

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10 

11 

Figure 8.9. Chlorophyll a profiles of pit lakes (a) Ramp 5a, (b) Ramp 8, (c) GAEE, (d) GAEW from May 2005 to 

August 2007. 

78 

0 

5 

10 

15 

20 

25 

30 

35 

40 

45 

50 

55 

60 

65

8.3.4 Metal/metalloid concentrations 

79 


There was greater variability in metal/metalloid concentrations than expected (Figure 8.10). 

Aluminium concentrations were generally stable over time, with highest concentrations of 

around ca.900 mg L -1 in Ramp 5a and lowest to ca.40 mg L -1 in Ramp 8 with GAEE/GAEW 

intermediate at 190 mg L -1. Ramp 5a surface waters Al concentration dropped drastically to 

ca.190 mg L -1 following heavy Wet season rains in January 2007, before slowly increasing 

again. GAEE and GAEW surface Al concentrations also decreased to 170 mg L -1 and 

120 mg L -1 respectively at this time, before also beginning to return to previous 

concentrations, albeit interrupted by further rains in June 2007. Although pit lake bottom 

waters remained largely unchanged over the monitoring period in Control lakes, GAEW 

bottom waters also decreased following 2006–2007 Wet season rains to 120 mg L -1 at the end 

of monitoring in July 2007. 

Conversely, B (boron) concentrations were very inconsistent over time at ranging from below 

detection limit 0.05 mg L -1 to


Ramp 5a and GAEW surface waters, which dropped following heavy rains in early 2007, Cd 

concentrations of other lakes stayed relatively stable to the end of monitoring. 

Cobalt concentrations were also high in all pit lakes. They were highest in Ramp 5a at around 

5 mg L -1 , lowest at 0.9 mg L -1 in Ramp 8 and 2.5 mg L -1 in GAEE/GAEW. Cobalt 

concentrations held at very constant concentrations for all lakes, except for Ramp 5a and 

GAEW surface waters which both showed decreases to around 0.9 mg L -1 following heavy 

Wet season rains in early 2007 before beginning to increase again in the following Dry 

season. 

Starting at high concentrations of around 0.4 and 0.15 mg L -1 for Ramp 5a bottom and 

surface waters respectively, and GAEE/GAEW at around 0.06 mg L -1, these three pit lakes 

also held relatively constant Chromium (Cr) concentrations until early 2007. Wet season 

rains occurred whereupon Ramp 5a surface water concentrations dropped sharply to around 

0.04 mg L -1. GAEE/GAEW surface waters also saw weaker declines to similar 

concentrations. Although GAEE/GAEW bottoms waters did not show significant Cr 

concentration changes during this time, the second-half of 2007 saw declines in GAEW 

bottom waters to around of 0.03 mg L -1 . Ramp 8 monitoring Cr concentrations began below a 

detection limit of 0.01 mg L -1 , rose to 0.02 mg L -1 in bottoms waters and 0.03 mg L -1 in 

surface waters over the second half of 2005 and first half of 2006, to again largely falling 

below the detection limit after the end of the 2006 Dry season. 

Copper concentrations followed a very similar trend to Cr, starting at high concentrations of 

around 2.2 and 1.0 mg L -1 for Ramp 5a bottom and surface waters respectively, and 

GAEE/GAEW at around 0.03 mg L -1. Nevertheless, although Ramp 5a and GAEE/GAEW 

surface waters showed a similar decline following early 2007 Wet season rains, aside from an 

outlier peak, GAEW bottoms water showed a decline to below a detection limit of 0.05 mg L - 

1 for this analyte in the first-half of 2007. 

Albeit with concentrations around 2,000 and 1,000 mg L -1 for Ramp 5a bottom and surface 

waters respectively, and GAEE/GAEW around 1,000 mg L -1 , Fe also showed a similar trend 

as Cr. Relatively stable Fe concentrations existed early 2007 Wet season rains, then there 

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81 


were sudden decreases to 17 and 420 mg L -1 for Ramp 5a and GAEW surface waters. These 

decreases were then followed by increases to concentrations at July 2007 below those of 

initial monitoring concentrations for Ramp 5a waters and at the initial monitoring 

concentrations for other lakes. As with other solutes, Ramp 8 Fe concentrations were lower 

for Fe than other lakes, with bottom waters around 400 mg L -1 and surface waters around 

1,100 mg L -1 . 

Starting at concentrations of around 1,000–1,800 mg L -1 , Mg showed a different trend to the 

other alkali metal Ca, but similar to that of the heavy metals. Following early 2007 rains, Mg 

concentrations declined to 270 and 590 mg L -1 in Ramp 5a and GAEW surface waters 

respectively. All lakes and depths then increased back to initial monitoring concentrations 

over the 2007 Dry season, albeit with Ramp 5a surface waters the slowest to return. 

Manganese also showed a similar trend to other heavy metals, with initial monitoring Mn 

concentrations of around 100 mg L -1 for Ramp 5a bottom waters, 40 mg L -1 for both Ramp 8 

depths and 60 mg L -1 for other lakes and depths. Following early 2007 rains, Mn 

concentrations declined to 15 and 30 mg L -1 in Ramp 5a and GAEW surface waters 

respectively before then beginning to increase back again. 

Initial monitoring Na concentrations of all lakes were around 650 mg L -1 , except for Ramp 5a 

with bottom waters around 450 mg L -1 an surface waters around 1,000 mg L -1 . Following 

early 2007 rains, Na also showed declines to 160 and 270 mg L -1 in Ramp 5a and GAEW 

surface waters respectively. All lakes and depths then increased back to initial monitoring Na 

concentrations over the 2007 Dry season, albeit with Ramp 5a surface waters the slowest to 

return. Ramp 5a bottom waters had reached around 1,000 mg L -1 by the end of monitoring in 

July 2007. 

Nickel concentrations also followed the general heavy metal trend of high initial monitoring 

concentrations of around 8 and 12 mg L -1 for Ramp 5a surface and bottom waters 

respectively, 2 mg L -1 for both Ramp 8 depths and 5 mg L -1 for GAEE/GAEW lakes and 

depths. Following early 2007 rains, Ni concentrations declined to 1.7 mg L -1 in both Ramp 5a


and GAEW surface waters before then beginning to increase back to around 5 mg L -1 in 

Ramp 5a surface waters. Both GAEW water depths remain below initial monitoring 

concentrations at around 3 mg L -1. 

High Pb (lead) concentrations in all lakes of around 3 mg L -1 decreased to below a detection 

limit of 0.1 mg L -1 from late Dry season 2005 to late Dry season 2006. Following early 2007 

rains, Pb concentrations declined to 1 and 0.25 mg L -1 in Ramp 5a and GAEW surface waters 

respectively before then beginning to increase back to around previous concentration levels 

by July 2007. 

Sulfur concentrations were at extremely high concentrations of 6,000 mg L -1 for Ramp 5a 

bottom waters and 3,500 mg L -1 for all other lakes and depths until the beginning of 2007 

Wet season rains. Sulfur concentrations then increased by 2–3 times for a few months, until 

then decreasing back to initial concentrations. At the end of July 2007 monitoring, Ramp 5a 

and GAEW surface waters continued to decrease past previous concentrations to around 

2,000 mg L -1 . 

Zinc concentrations were high around 30 and 40 mg L -1 for Ramp 5a surface and bottom 

waters respectively, 6 mg L -1 for both Ramp 8 depths and 15 mg L -1 for GAEE/GAEW lakes 

and depths. Following early 2007 rains, Zn concentrations declined to 7 mg L -1 and 5 mg L -1 

in Ramp 5a and GAEW surface waters respectively before then beginning to increase back to 

initial monitoring concentrations. However, following early 2007 rains, GAEW bottom water 

Zn concentrations declined erratically to a July 2007 concentration of around 1 mg L -1. 

82

83 

McCullough, Lund and May (2008)


84

85 

McCullough, Lund and May (2008)


Figure 8.10. Changes in metal concentrations of Collinsville pit lakes and layers from April 2005 to August 2007. Note log(10) y-axis scale. 

86

87 


At the project completion in July 2007, aside from Ca, most mean water column metal 

concentrations in GAEW were different to that of GAEE (Table 8.3). Some mean GAEW 

metals’ concentrations such as Cd, Cr, Cu Fe and Pb were up to 10–20% higher than in the 

control lake. Potassium was 5,600% higher concentration. Nevertheless, Al, Ca, Co, Mg, Mn, 

Na, Ni and S and Zn all showed decreases in concentration of around 1–30%. 

Table 8.3. Mean GAEW metal concentrations relative to mean GAEE concentrations at July 2007 (mg L -1 ). 

Maximum concentrations indicate concentrations if all dosed mass were in solution. 

Solute Al Ca Cd Co Cr Cu Fe K Mg Mn Na Ni Pb S Zn 

GAEE 200 240 0.07 2.5 0.06 0.34 1,070 0.5 850 63 610 4.7 2.6 3,400 15 

GAEW 140 230 0.08 1.8 0.07 0.42 1,220 28 790 57 550 3.1 2.8 3,200 

Maximum 

concentrations 

%Change 69 97 106 72 120 120 110 5,600 92 89 90 67 110 94 43 

8.3.5 Nutrients 

Total nitrogen levels of all control lakes were generally less than 10 mg L -1 although 

hypolimnion waters of Ramp 5A were generally highest and displayed the highest 

concentration of 22 mg L -1 in August 2006 (Figure 8.11). Total nitrogen levels increased in 

the hypolimnion of GAEE only, both during and after dosing, to around five times (TN, 

46 mg L -1 ) background levels of total nitrogen. Rapid increases in both TN concentration to 

around 46 mg L -1 and TP to around 10 mg L -1 occurred after filling had ceased in January 

2007. Peculiarly, in July 2007 TN displayed a sudden decrease from around 45 mg L -1 to 

7 mg L -1 and TP concentrations dropped from 10 mg L -1 to around 5.8 mg L -1 . Concentrations 

of TN nevertheless rose over the next quarter to around 42 mg L -1 . 

Total phosphorus of control lakes was generally under 1 mg L -1 (Figure 8.11). However, 

hypolimnion waters of Ramp 5A were often elevated in TP, cycling every 4–5 months from 

1 mg L -1 to a high of 6 mg L -1 in September 2006. Epilimnion waters of Ramp 5A were 

generally under 1 mg L -1 although they suddenly spiked to 4 mg L -1 in April 2007. Total


phosphorus levels rapidly increased soon after filling began in August 2006, first in the 

epilimnion of GAEE to 9 mg L -1 before increasing in the hypolimnion to 10 mg L -1 around a 

month later. Total phosphorus epilimnion concentrations then declined in February 2007 and 

fluctuated around 4 mg L -1 afterward. GAEE hypolimnion waters also declined to 6 mg L -1 

around 3 months later, before spiking back to 10 mg L -1 the next month and then declining 

again. 

Gilvin of all control lakes displayed very high levels likely to be caused by confounding by 

the high dissolved iron concentrations of the AMD matrix (Figure 8.11). Nevertheless, 

GAEW epilimnion gilvin did appear to increase after organic additions in August 2006, 

peaking at 187 m -1 in June 2007, before declining to 110 m -1 in July 2007. 

The large additions of nutrient-rich organic material to GAEW resulted in greatly elevated 

concentrations of macro-nutrients (Table 8.4). However, there was a large difference between 

mean GAEW measured concentrations and maximum potential nutrients concentrations 

(assumed with a model of nutrients completely mixed but behaving conservatively). 

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89 


Figure 8.11. Changes in (a) total nitrogen (TN), (b) total phosphorus (TP) and (c) dissolved organic carbon (as 

gilvin440) concentrations of Collinsville pit lakes and depths from April 2005 to August 2007.


Table 8.4. Mean depth GAEW nutrient concentrations compared to mean GAEE concentrations at July 2007 

(mg L -1) . 

8.3.6 Groundwater interactions 

TN TP 

GAEE 0.50 0.47 

GAEW 21 4.5 

GAEW maximum concentrations 

Factor of change 43 9.6 

Three water level data outliers were removed from the dataset; the first point for bore GAR01 

as it was assumed the bore was still stabilising after establishment, and a data point from 

GAR01 and GAEE/GAEW as for a single month they deviated grossly from surrounding data 

dates and monitoring points. 

Throughout the 46 month sampling period, groundwater levels were consistently above 

GAEE/GAEW water surface levels (Figure 8.12). Groundwater height also greatly increased 

generally following heavy rainfall events in early 2007. Groundwater height was greatest in 

GAR01 to the west of GAEE/GAEW, and lowest to the south in GAR03 with GAR02 depth 

slightly greater than GAR03. 

pH was circum-neutral across all monitoring bores at all times. Although groundwater pH 

fluctuated from 0.5–1.0 units between sampling months, there may have been a trend of 

decreasing pH in the three years of monitoring (Figure 8.13a). Groundwater specific 

conductance was highest in GAR01 at ca.13 mS cm -1 and lowest in GAR03 at ca.2 mS cm -1 

with GAR02 intermediate ca.10 mS cm -1. Groundwater specific conductance fluctuated very 

little in the southern monitoring bore GAR03. Groundwater at GAR03 did not respond in the 

same way as the more northern bores where there were troughs in specific conductance in 

April 2005 and October 2006 and a large spike in around July 2007 (Figure 8.13c). 

Groundwater solute concentration in all monitoring bores were dominated by SO 4 , Cl, Ca and 

Na (Figure 8.14). The proportion of major solutes were similar in the northern-most bores 

GAR01 and GAR02, however calcium represented only 7% of these total solute 

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91 


concentrations in the southern-most bore GAR03. Reflecting relative specific conductance 

values, major solute concentrations appeared to be fairly stable over time. Groundwater major 

solute concentrations were highest in eastern-most GAR02 and lowest in southern-most 

GAR03. 

Lower concentrations of Al, As, B, Co, Cr, Cu, Li, Mn, Mo, Ni, Se, U and Zn, were also 

detected in groundwaters, although Be, Bi, Sb and V were all below detection limits. Iron 

was the highest concentration solute for all groundwaters, and then Mn although southern- 

most GAR03 had higher concentrations of F than Mn (Figure 8.15). Although more variable 

than major solutes, relative and absolute concentrations of minor solutes still appeared to be 

fairly stable over time. Groundwater minor solute concentrations also reflected relative 

specific conductance values, being highest in eastern-most GAR02 and lowest in southern- 

most GAR03. 

Figure 8.12. Relative GAEE/GAEW groundwater and water surface levels between July 2004 to October 2007.


Figure 8.13. GAEE/GAEW groundwater monitoring bore water a) pH, b) acidity and, c) specific conductance 

between August 2004 to October 2007. 

92 

b)

Concentration (mg L -1) 

93 


Figure 8.14. Changes in major solute concentrations in monitoring bores a) GAR01, b) GAR02, and c) GAR03 

between November 2005 and July 2007.


Concentration (mg L -1) 

Figure 8.15. Changes in minor solute concentrations with mean values


95 


Remediation treatment of GAEW after only 6 months still appears to be in the early stages of 

development, however some trends have already become evident. Depth and duration of 

stratification is an important decider in the efficacy and viability of sulfate reduction as a 

remediation measure in any pit lake (Martin et al., 2003; McNee et al., 2003). Generally, 

CCP pit lakes appear to thermally stratify in the late Dry season although this stratification is 

broken when Wet season wind and rains arrive. Consequently, water column stratification is 

expected to occur in the treatment pit lake for at least 6 months of the year. Although heavy 

cyclonic rainfall events, as seen in early 2007 may destabilise thermal stratification earlier 

than usual, the halocline that forms as a result of this low conductivity water overlaying the 

denser high conductivity pit lake water may be more stable and effective at preventing 

mixing and consequently reduced metal oxidation than thermal stratification during warmer 

months (Wen et al., 2006). As shown by 2006 data, the unpredictable climate of this site 

makes prediction of remediation performance difficult as heavy rains and mixing events can 

still occur at any time of the year. 

Pit lakes do not appear to be thermally stratified over the cooler Dry season. Inversion of the 

thermocline during cooler months of the mid-Dry season are not likely to be due to 

exothermic re-oxidation of reduced sulfides in the hypolimnion as has been observed with 

cool-temperate AMD pit lakes (Gammons & Duaine, 2006) as the hypolimnion ORP did not 

markedly change at this time under the continuing influence of the halocline. Rather, it this 

inverted thermocline is more likely due to exothermic biological activity of anaerobic 

bacteria breaking down the rich the organic material. 

Low ORP groundwater appears to influence hypolimnions of Ramp 5a, and Ramp 8 in 

particular, it does not appear to be significantly intruding into GAEE or GAEW. The low- 

ORP cycles of Ramp 8 in mid-Dry season and mid-Wet season may be due to low pit lake 

levels due to evaporation, and high groundwater levels due to precipitation in the Dry and 

Wet seasons respectively. The water level differences would lead to hydraulic head in the 

favour of groundwater intrusion during these times. Similarly, although low ORP rainwater 

improved water quality for some time across Ramp 5a, the size of GAEE and GAEW appears


to have reduced the impact of the rainfall on ORP in these lakes. Nevertheless, catchment 

area the decreasing ORP in GAEW hypolimnion waters appears to be due to sulfate reductive 

processes in the lake’s sediment. These reductive process may also have been facilitated by 

halocline formation early in the 2006–2007 Wet season. 

Other control pit lakes have seen a slight decrease in epilimnion pH due to cyclonic rain 

waters, and Ramp 8 has seen a significant increase in hypolimnion waters due to groundwater 

ingress. However, unlike the very similar GAEE, GAEW demonstrated an increasing pH 

developing in the hypolimnion since late 2006, only a few months after organic dosing began. 

This pH increase is indicative of alkalinity-producing sulfate reduction in these low ORP 

waters. 

Not surprisingly, phytoplankton biomass is generally very low in the low nutrient 

environment of CCP pit lakes. Nevertheless, the addition of large amounts of macronutrients 

(carbon and phosphorus in particular) appears to have stimulated phytoplankton growth in 

GAEW a few months after organic dosing began. It is conceivable that benthic microbial 

communities may be supporting phytoplankton communities through reducing iron and 

aluminium bound P, effectively remobilising it, making it again available for absorption by 

phytoplankton (Kleeberg & Grüneberg, 2005). In turn, phytoplankton may be supporting 

sulfate reduction by fixing carbon in photosynthesis with dead algal cells then falling to the 

benthos where this carbon is made available to bacterial sulfate-reducers (Nixdorf et al., 

2005). 

The sudden mid-2007 increase in surface water concentrations of metals to previous 

concentrations, including conservative ions such as Na may be due to evapo-concentration of 

epilimnions previously diluted by the heavy early 2007 Wet season rains. Although halocline 

epilimnion waters are lower EC, the slightly lower EC of GAEW bottom waters is probably 

primarily due to precipitation of iron and sulfate as monosulfides as indicated by decreases in 

S concentrations. Other, lower solubility metals such as Zn may also have precipitated as 

(oxy)hydroxides in this higher pH environment. The low EC that has been maintained in 

surface waters of well-mixed GAEW in July 2007 may similarly be due to increased algal 

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activity adsorbing some solutes (in particular nitrogen compounds) and again forming a 

higher pH local environment precipitating some metals by producing alkalinity during nitrate 

assimilation. 

Consequently, changes in metal/metalloid concentrations in the epilimnion can partially be 

explained through dilution during heavy rainfall events, and in the hypolimnion by 

immobilisation as sulfide and hydroxides as a result of sulfate reduction processes. 

Notwithstanding the effects of these processes, there was greater variability in hypolimnion 

metal/metalloid concentrations than expected. This may be due to subtle but significant 

differences in sampling height above the benthos at different sampling times. As the benthos 

is neared, a steep ORP gradient becomes apparent, and this is likely to correspond to a steep 

solute gradient for non-conservative species. 

Fe(III) reduction has been observed as the initial step of microbial sulfate reduction (Wendt- 

Potthoff et al., 2002). Although there have not yet been significant reductions in total 

dissolved Fe or S concentrations indicative of sulfate reduction, this may be due to Fe(III) 

reduction and consequent mobilisation showing an increase in total dissolved Fe 

concentration, prior to iron sulfide formation. 

The initial high concentration of chlorophyll in GAE at the beginning of monitoring in mid- 

2005 is difficult to explain as water quality throughout monitoring was extremely poor for 

algal growth. Nevertheless, as found with other studies eutrophying acid pit lakes (Lessmann 

et al., 2000; Fyson et al., 2006), chlorophyll concentrations, an indicator of phytoplankton 

biomass, have been shown to increase quickly following additions of limiting carbon and 

phosphorus nutrients. Consequently, the algal blooms of GAEW months after dosing are not 

unexpected. 

Although a northern monitoring bore is absent due to coal outcropping there, local 

groundwater movement appears to be from north to south toward nearby Corduroy Creek. 

Nevertheless, the consistently greater height of groundwater around the pit lake indicating 

that the pit lake is a local groundwater sink, with contamination of downstream groundwater


by leachate from this pit unlikely. Specific conductance appears to increase from west to east 

across GAE, however not across the expected flow direction of north to south implying a 

localised source of groundwater contamination immediately north of GAE. The elevated iron 

in the presence of marine salts implies that AMD evolution may be occurring in the presence 

of buffered saline groundwaters in this region, possibly originating from prehistoric saline 

intrusions. These waters, although circum-neutral, are likely introducing acidity and salinity 

to the GAE pit lakes. 

98

9 Conclusions 

99 


For long-term remediation purposes, organic matter additions are expected to be required at 

an ongoing, albeit lower, dosing rate. The use of refractory organic forms such as woody 

green waste, may also be advantageous to long-term remediation efforts, in that they will 

continue to degrade into more labile fractions available for sulfate reducers over long periods 

of time. Nevertheless, whilst denitrification is another alkalinity-generating process (Abril & 

Frankignoulle, 2001) that will remove excess TN over time, over-dosing of organic matter or 

changes in availability of P may lead to eutrophication (Yokum et al., 1997). Depending on 

final desired water quality and the location, eutrophication might be of concern (McCullough 

et al. (2008). It is unlikely, given the high iron levels that eutrophication of GAEW is likely 

in the short term to medium term, however as the rate of remediation is likely to increase as 

we move from extreme to more circum-neutral pH caution is recommended with regards to 

the quantities and nature of further organic matter additions. Nevertheless, eutrophied water 

bodies have an extensive research history and ‘toolbox’ for management, and will likely 

present more benefit to the environment and communities than the otherwise acidic pit lakes 

do. 

Depth and duration of stratification will remain an important decider in the efficacy and 

viability of sulfate reduction as a remediation measure in any pit lake (Martin et al., 2003; 

McNee et al., 2003). Water column stratification is expected to remain strong in this 

remediated pit lake over late Dry seasons due to the ambient hot tropical climate. In much the 

same way as straw has been postulated to change mixing in other studies (Koschorreck et al., 

2002) greenwaste is likely to act as a mixing barrier for the water body directly above the 

sediment surface. Coupled with a high biochemical oxygen demand from continuing organic 

decomposition and carbon diagenesis, the hypolimnion should remain sufficiently anaerobic. 


(Berrow & Webber, 1971). Both the Collinsville sewage displayed high concentrations for 

many heavy metals. However, these high heavy metals are unlikely to be biologically 

available, as if mobilised are likely to be precipitated as metal sulfides.


Further reduction in biological availability of metals and metalloids is also likely to occur 

through formation of complexes with organic chelators present as components of the 

refractory green waste, such as organic acids (Tipping & Hurley, 1992). In this respect, green 

waste may have an advantage of providing recalcitrant organic substances such as humic and 

fulvic acids, with which heavy metals may directly complex to. Ligand formation between 

heavy metals and refractory organics would further remove these toxic components from 

biological availability, albeit at a likely reduced capacity to that of sulfate reduction processes 

(Brown Jr., 2001). 

The early initiation of alkalinity production in green waste and green waste and sewage in 

only 37 days is likely to be due to a combination of two important factors unique to this 

study. Firstly, most other research published to date have occurred in cool temperate areas of 

Europe and North America e.g., (Gammons et al., 2000; Küsel & Dorsch, 2000; Küsel et al., 

2001; Benner et al., 2002; Tostche et al., 2003; Frömmichen et al., 2004). The higher 

temperatures experienced in Collinsville, even during the middle of the Dry season (20– 

28 o C), and mimicked in the constant-temperature laboratory are likely to have exponentially 

increased biochemical rates of carbon diagenesis and metabolism. 

Secondly, the use of largely fresh green waste as opposed to refractory organic substrates 

such as straw (Frömmichen et al., 2003; Frömmichen et al., 2004), rye grass (Harris & 

Ragusa, 2000, 2001), etc., distinguishes this research from many others in the published 

literature. The greater labile fraction of organic material available in this fresher material may 

have contributed directly to electron donors for sulfate reduction. For example, chlorophyll 

was seen to be leached from the green waste and it is likely that highly labile sap sugars 

would have leached also. For long-term remediation purposes, organic matter additions are 

expected to be required at an ongoing, albeit lower, dosing rate. In this way, the use of 

refractory organic forms such as woody green waste, may be advantageous to long-term 

remediation efforts, in that they will continue to degrade into more labile fractions available 

for sulfate reducers over long periods of time (Place et al., 2006). Consequently, the use of 

fresh sewage to acid pit lakes, with or without complementary additions of green waste, may 

prove to be a novel practicable remediation strategy for AMD issues in remote mining 

locations. Indeed, Gusek (2002) even suggested injecting sewage sludge into mine shafts and 

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101 


adits to remove and prevent production of acidity from these sources. For long-term 

remediation purposes, organic matter additions are expected to be required at an ongoing, 

albeit lower, dosing rate. The use of refractory organic forms such as woody green waste, 

may also be advantageous to long-term remediation efforts, in that they will continue to 

degrade into more labile fractions available for sulfate reducers over long periods of time. 

The ongoing monitoring of the field study component of this project is expected to reveal 

these sustainability considerations over these long treatment durations. 

This study has demonstrated promise for remediation of AMD water pH and high metal 

concentrations via sulfate reduction with green waste and sewage as organic substrates. This 

study also illustrates that corrective measures for acidic mining lakes also greatly differ from 

those in use for eutrophication control (Klapper, 2003). For example, average species 

number, bird numbers and biomass have been found to positively correlated with lake trophic 

status (Hall et al., 2006), demonstrating greater potential wildlife value for a eutrophic than 

an ultra-oligotrophic mining lake. 

Best practice AMD pit lake treatment is likely to be required to be made on a case-by-case 

basis where potential pit lake end uses and availability of organic substrates have been 

identified and a remediation strategy has consequently been constructed. Consequently 

researchers, consultants and regulatory agencies need to maintain an open-mind to treatment 

solution options for these unique environmental issues. Part of the challenge for both mining 

companies and regulatory waterbody managers with this assessment will be the ability to 

think laterally as to what the current values of the pit lake are, what treatments may be 

feasible, and what type of pit lake (water quality and quantity) is desirable for either social, 

environmental or other enduses (Doupé & Lymbery, 2005; McCullough & Lund, 2006).


10 Acknowledgements 

Thanks to Collinsville Coal Project (Xstrata Coal Queensland Pty Ltd) for logistical support 

and lease access. Thanks to Bowen Shire for in-kind support with materials and transport. 

Thanks to Queensland Environmental Protection Authority for consideration of the field 

project component. Thanks also to Joseph Steenbergen and Carlieke te Beest of Hogeschool 

Zeeland for laboratory assistance, to Gary Ogden for some of the chemical analyses, and to 

Tim Walker and Carl Wallis for field assistance. This project was made possible by 

Australian Coal Association Research Council grant C14052. 

102

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112

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12 Appendix: Publications and Presentations 

arising from ACARP Project C14052 

Lund, M. A. & McCullough, C. D. (2004). Mine Lakes - More than just acidic, toxic scars on 

the landscape. In, Australian Society for Limnology 43rd Annual Congress. Adelaide, 

Australia 29th November - 3rd December, 2004. 

May, J.M., McCullough, C. D. & Lund, M. A. (2007), “Using Bioremediation to Tackle Acid 

Mine Drainage Water at Collinsville Coal Mine, QLD", Presentation made at XCN 

Environment & Community Forum, Kirkton Park, NSW, 7th May 2007. 

May, J.M. (2006). "Using bioremediation to tackle acid mine drainage water", Presentation 

made to Retired Miners’ Association, Collinsville, QLD,18th July 2006. 

May, J.M.(2006). "Using bioremediation to tackle acid mine drainage water", Presentation 

made to CFMEU Collinsville Lodge Representatives, Collinsville, QLD, 20th July 

2006. 

May, J.M. (2005), "A case study aimed at assisting CCP to meet its XCQ Sustainable 

Development, Biodiversity, Water, Community & Rehab Targets: Remediation of acid 

coal mine voids using biological processes and organic material", Presentation made at 

XCQ Environment & Community Meeting, Abbot Point, QLD, 17th August 2005. 

May, J.M., McCullough, C. D. & Lund, M. A. (2005). "ACARP Project C14052: 

Remediation of acid coal mine voids using biological processes and organic material", 

Presentation made at Central Queensland Mining Rehabilitation Group Meeting, Oaky 

Creek, QLD, 25th October 2005. 

McCullough, C. D. (2007). Approaches to remediation of acid mine drainage water in pit 

lakes. International Journal of Mining, Reclamation and Environment. 21 DOI: 

10.1080/17480930701350127. 

McCullough, C. D.; Hunt, D. & Evans, L. H. (in press). Social, Economic, and Ecological 

End Uses – Incentives, regulatory requirements and planning required to develop 

successful beneficial end uses. In Workbook of Technologies for the Management of 

Metal Mine and Metallurgical Process Drainage, Castendyk, D.; Eary, T. & Park, B. 

(eds.) Society for Mining Engineering (SME), Kentucky, USA, 

McCullough, C. D. & Lund, M. A. (2006). Pit lakes: benefit or bane to companies, 

communities and the environment? Proceedings of the Goldfields Environmental 

Management Group Workshop on Environmental Management 2006. Kalgoorlie, 

Australia 24th - 26th May. 12p. 

McCullough, C. D.; Lund, M. A. & May, J. M. (2005). The addition of green waste and 

municipal sewage to a tropical acid pit lake, a novel approach to remediation; or, 'Why 

we filled a pit lake with dead plants and poo'. Proceedings of Australian Society for 

Limnology 44th Annual Congress. Hobart, Australia 28th November - 2nd December, 

2004. 


sewage and green waste for full-scale remediation of an acid coal pit lake, in semi-arid 

tropical Australia. Proceedings of the 7th International Conference on Acid Rock 

Drainage (ICARD). St Louis, Massachusetts, USA. 1,177-1,197.


McCullough, C. D. & Lund, M. A. (2006). Pit lake sustainability; what is it, and how do I get 

it? Proceedings of the 2006 Water in Mining conference. Brisbane, Australia 24th - 

26th November, Australasian Institute of Mining & Metallurgy. 323-330. 

McCullough, C. D. & Lund, M. A. (2006). Opportunities for sustainable mining pit lakes in 

Australia. Mine Water and the Environment. 25(4): 220-226. 

McCullough, C. D.; Lund, M. A. & May, J. (2008). Field scale demonstration of the potential 

for sewage to remediate acidic mine waters. Mine Water and the Environment. 27(1) 

DOI 10.1007/s10230-007-0028-y. 

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