Chemosphere 103 (2014) 172–180

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Chemosphere

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Metal speciation and potential bioavailability changes during dischargeand neutralisation of acidic drainage water

0045-6535/$ - see front matter Crown Copyright � 2013 Published by Elsevier Ltd. All rights reserved.http://dx.doi.org/10.1016/j.chemosphere.2013.11.059

⇑ Corresponding author. Tel.: +61 2 9710 6807.E-mail address: stuart.simpson@csiro.au (S.L. Simpson).

Stuart L. Simpson a,⇑, Christopher R. Vardanega a,b, Chad Jarolimek a, Dianne F. Jolley b, Brad M. Angel a,Luke M. Mosley c

a Centre for Environmental Contaminants Research, CSIRO Land and Water, Locked Bag 2007, Kirrawee, NSW 2232, Australiab School of Chemistry, University of Wollongong, New South Wales, Australiac Environment Protection Authority South Australia, GPO Box 2607, Adelaide, SA 5001, Australia

h i g h l i g h t s

� Discharge of acid drainage from farm irrigation areas represents a risk to ecosystem health.� Rapid precipitation of Al and Fe increases removal of other metals from dissolved phase.� Many dissolved metals in labile and potential bioavailability forms.� Similar factors controlling the dissolved concentrations (pH, dilution and mixing time).� Water quality guideline exceedance is unlikely for drainage waters dilution to 1%.

a r t i c l e i n f o

Article history:Received 12 August 2013Received in revised form 21 November 2013Accepted 22 November 2013Available online 18 December 2013

Keywords:Acid sulfate soilsWater quality guidelinesEcotoxicologyRisk assessmentDairy farmMurray River

a b s t r a c t

The discharge of acid drainage from the farm irrigation areas to the Murray River in South Australiarepresents a potential risk to water quality. The drainage waters have low pH (2.9–5.7), high acidity(up to 1190 mg L�1 CaCO3), high dissolved organic carbon (10–40 mg L�1), and high dissolved Al, Co, Niand Zn (up to 55, 1.25, 1.30 and 1.10 mg L�1, respectively) that represent the greatest concern relativeto water quality guidelines (WQGs). To provide information on bioavailability, changes in metal specia-tion were assessed during mixing experiments using filtration (colloidal metals) and Chelex-lability (freemetal ions and weak inorganic metal complexes) methods. Following mixing of drainage and river water,much of the dissolved aluminium and iron precipitated. The concentrations of other metals generallydecreased conservatively in proportion to the dilution initially, but longer mixing periods causedincreased precipitation or adsorption to particulate phases. Dissolved Co, Mn and Zn were typically95–100% present in Chelex-labile forms, whereas 40–70% of the dissolved nickel was Chelex-labile andthe remaining non-labile fraction of dissolved nickel was associated with fine colloids or complexed byorganic ligands that increased with time. Despite the different kinetics of precipitation, adsorption andcomplexation reactions, the dissolved metal concentrations were generally highly correlated for thepooled data sets, indicating that the major factors controlling the concentrations were similar for eachmetal (pH, dilution, and time following mixing). For dilutions of the drainage waters of less than 1% withMurray River water, none of the metals should exceed the WQGs. However, the high concentrations ofmetals associated with fine precipitates within the receiving waters may represent a risk to some aquaticorganisms.

Crown Copyright � 2013 Published by Elsevier Ltd. All rights reserved.

1. Introduction

Submerged soils and sediments are frequently observed toaccumulate pyritic (FeS2) phases as a result of the sulfate reductionthat occurs naturally through microbial respiration of organic car-bon (Dent and Pons, 1995). When undisturbed and covered with

water, the pyrite poses little or no threat of acidification, howeverconsiderable oxidation and generation of acidity can occur whenpyrite is exposed to the air (Bronswijk et al., 1993; Dent and Pons,1995). The rewetting of oxidised acid sulfate soils may release sig-nificant quantities of metals to associated water (Cook et al., 2000;Simpson et al., 2010; Nystrand et al., 2012).

Drought conditions and long-term low inflows in the Murray-Darling Basin system from 2006 to 2010 led to unprecedentedlow water levels in the lower river reaches in South Australia

S.L. Simpson et al. / Chemosphere 103 (2014) 172–180 173

(Mosley et al., 2012). During April 2009, the river water level fell to�1.75 m below average long-term levels and exposed large areasof soils to prolonged drying and desiccation, which resulted inthe oxidation and acidification of previously submerged acid sul-fate soils in the Lower Murray Reclaimed Irrigation Area (LMRIA).Since late 2010, water levels have recovered in the Lower Murray,irrigation has recommenced in some areas, and the drought-af-fected soils have been rewetted. Acidic drainage waters with pH2–5 are now being returned (via large drainage pumps) to the Mur-ray River from a range of irrigation areas in the LMRIA (EPA, 2013).The low pH and high concentrations of soluble and potentiallytoxic metals within the drainage waters posed risks to aquatic eco-systems and drinking water off-takes in the river.

The risk posed by the acid drainage water discharges to the Mur-ray River ecosystem will be strongly influenced by the form and fateof the metals and metalloids during mixing (Gundersen and Stein-nes, 2003; Balistrieri et al., 2007; Nordstrom, 2011; Cresswell et al.,2013). Upon discharge and mixing with the river water alkalinity,pH neutralisation reactions can be expected to result in the floccu-lation and precipitation of a considerable portion of many of themetals, and adsorption of other metals or metalloids onto thesenew solid phases (Gundersen and Steinnes, 2003; Balistrieri et al.,2007; Lee and Faure, 2007; Schemel et al., 2007). Previous studieshave observed a high degree of colloid formation, e.g. 75% for alu-minium (Gundersen and Steinnes, 2003; Schemel et al., 2007) and�40% for iron (Schemel et al., 2007) following mixing of acid-drain-age waters with ambient waters. However, those studies were foracid-mine drainage, where the dissolved organic carbon (DOC) con-centrations were much lower (0.2–2 mg L�1) than the 10–40 mg L�1 range in the LMRIA acid drainage. It was unclear whethermetals in the LMRIA acid drainage would show similar behaviourand potential bioavailability compared to acid-mine drainage sites.

The speciation of the metals in the dissolved phase will influ-ence their bioavailability and the hazard posed. This may be med-iated by other cations that compete for metal binding sites, bymetal complexation by organic ligands, and by metals in colloidalforms (van Dam et al., 2008; Fortin et al., 2010; Gandhi et al.,2011). Australia’s water quality guidelines (WQGs) for the protec-tion of aquatic ecosystem health were derived using species sensi-tivity distributions of chronic no observed effects concentrations(NOECs) or 10% chronic effect concentrations (EC10’s) derived pre-dominantly from laboratory-based bioassays on solutions of met-als in highly labile and bioavailable forms, e.g. predominantly asfree metal ions (e.g. M2+) or weak and labile inorganic metal com-plexes, with negligible strong metal–ligand complexes that arekinetically non-labile such as metal complexes with humic or ful-vic acids (ANZECC/ARMCANZ, 2000). In the application of theWQGs, when there is evidence that, in the test waters, a significantportion of the total dissolved concentration is present in forms thatare not considered bioavailable (e.g. non-labile strong metal–li-gand complexes that occur with humic and fulvic acids and or col-loid-associated metals), it is appropriate to make the WQGcomparison with just the labile fraction (determined either bymeasurement or modelling). In the case of acidic drainage watersentering the Murray River, the proportions of dissolved metals inbioavailable forms are expected to vary both spatially and tempo-rally. Under acid conditions, free metal ion concentrations arehigher, but under more neutral conditions, complexation becomesimportant as does the association with precipitating iron oxyhy-droxides, reducing bioavailability.

The persistence and severity of the acid drainage in the LMRIAhas created considerable challenges to management agencies. Toaddress these risks, this study investigated the fate and forms ofmetals and metalloids following dilution and neutralisation ofthe drainage water with river water. The changes in metal concen-trations and speciation (e.g. complexed or colloidal forms) as acidic

drainage waters undergo neutralisation with river water wereinvestigated and the potential for ecological effects assessed.

2. Material and methods

2.1. Study area

The LMRIA comprises approximately 5200 hectares of flood irri-gated agriculture protected by a levee bank system on the formerfloodplain of the Murray River in South Australia, between thetownships of Mannum and Wellington (Fig. 1). The irrigation chan-nels are typically 1.0–1.5 m below the normal river pool level(+0.75 m Australian Height Datum), enabling gravity fed flood irri-gation. The drainage waters within the channels are returned tothe river using large pumps (Mosley and Fleming, 2010) (see con-ceptual model in Fig. S1 of the Supplementary information). Theregion contains important aquatic ecosystems and is immediatelyupstream of Coorong-Lower Lakes system which is collectivelyrecognised as one of Australia’s most significant ecological assetsand a wetland of international importance (Ramsar-listed).

2.2. Sample collection

The LMRIA sites chosen for analysis were Toora, Mobilong, LongFlat, Woods Point and Wellington (at the ends of Jervois) (Fig. 1).Murray River water was collected from the Thiele Reserve in Mur-ray Bridge to be used as the diluent in for drainage water dilutions.The Supplementary information section provides GPS coordinatesof the sites (Table S1). The water pH, temperature, specific electri-cal conductivity (EC), redox potential (ORP), dissolved oxygen (DO)concentration, total dissolved solids (TDS) were measured at thetime of sampling using a calibrated water quality probe (YSI556). Alkalinity and acidity (mg L�1 CaCO3) were measured in thefield using a test kit (HACH Model AL/AC-DT, mg L�1 CaCO3 canbe converted to meq L�1 by dividing by 50).

Water samples for analyses of trace metals and metalloids werecollected from all locations on March 22, 2012, using strict proto-cols to avoid sample contamination (Ahlers et al., 1990). Deionisedwater (18 MX cm, Milli-Q, Millipore) and high-purity acids (Trace-pur, Merck, Darmstadt, Germany) were used for washing of bottles(Nalgene, LDPE) and filters and for sample acidification. All plasticware was acid-washed by soaking for >24 h in 10% v/v HNO3, thenrinsing with copious amounts of deionised water (Milli-Q) beforedrying in a laminar-flow cabinet (Clyde-Apac, HWS Series) priorto use.

All samples were collected in acid-washed low-density polyeth-ylene (Nalgene) bottles (1 L bottles and 5 L carboys). Field dupli-cates and field blanks were included as part of the watersampling program. Following collection, the water samples weretransported to the laboratory in ice-filled cooler boxes and thenstored at 4 �C. Water filtration through 0.45-lm membrane filterswas undertaken in a Class-100 cleanroom laboratory as describedin Cresswell et al. (2013). This involved the use of acid-washed0.45 lm membrane filters (Millipore HA, mixed cellulose esters)and polycarbonate filter unit (Sartorius). Samples for dissolvedmetals or metalloid analyses were preserved in 0.2% HNO3, dis-solved organic carbon (DOC) preserved in 0.4% H2SO4, and samplesfor alkalinity, major ions and labile metal analyses contained nopreservatives. All sample bottles and containers were stored inplastic bags and refrigerated in the dark.

2.3. General methods

All glass and plasticware for analyses were usually new andwere cleaned by soaking in 10% (v/v) HNO3 (BDH, Analytical

Fig. 1. Map of LMRIA showing the drainage discharge sites (stars) and township (solid circles). The five sites where acid drainage water was sampled are labelled.

174 S.L. Simpson et al. / Chemosphere 103 (2014) 172–180

Reagent grade) for a minimum of 24 h, followed by thorough rins-ing with deionized water (Milli-Q, 18 MX cm). All chemicals wereanalytical reagent grade or equivalent analytical purity. In thelaboratory, measurements of dissolved oxygen were made usingan Oxi 196 WTW meter (Weilheim, Germany) with an oxygen elec-trode (EO96, WTW), conductivity using a LF 320 WTW conductiv-ity meter (Weilheim, Germany) and electrode (TetraCon 325,

WTW), and pH using a WTW meter equipped with a pH probe(Orion sure-flow combination pH 9165BN) and calibrated againstpH 4.0 and 7.0 buffers (Orion Pacific, Sydney, NSW, Australia)according to manufacturer’s instructions.

Acidified water samples were analysed for dissolved metals andmetalloids by inductively coupled plasma mass spectrometry(ICP–MS; Agilent 7500-CE) against standards prepared in acidified

S.L. Simpson et al. / Chemosphere 103 (2014) 172–180 175

Milli-Q, using an internal standard to correct for instrument driftand physical interferences. Due to the high ‘salt’ concentrationsin many of the waters, the samples were generally diluted withMilli-Q and standard addition methods used for dissolved metalsand metalloids analyses. Analyses of acidity/alkalinity (automatedtitration), chloride, sulfate, phosphate (using discrete analysers)and dissolved organic carbon (DOC) (high-temperature combus-tion method) were made following standard methods APHA(2005). Major ions were analysed by inductively coupled plasmaatomic emission spectrometry; ICP–AES; Varian 730-ES).

As part of the quality assurance (QA)/quality control (QC), anal-yses of filter and acid-digest blanks, replicates for 10% of samples,analyte sample spikes and certified reference materials (CRMs)were made. Replicates were within 10% and recoveries for spikesand the CRMs (TM 24.3 and TMDA 52.3, National Research Council,Canada) were within 85–115% of expected values. The limits ofreporting for the various methods were less than 1/10th of thelowest measured values.

2.4. Labile metal analyses

Information on the form and potential bioavailability of the fil-terable metals was derived from Chelex-labile metal measure-ments (Bowles et al., 2006). Greater metal lability usually resultsin greater metal bioavailability. Typical labile metals include freemetal ions and weak inorganic complexes while non-labile metalsinclude colloidal metals and strong metal complexes with organicligands such as naturally occurring humic acids. The concentra-tions of labile metals were determined according to the methoddescribed by Bowles et al. (2006). Briefly, sub-samples of filteredriver water were passed through a column containing Chelex-100resin at a precise and fast flow rate such that only free metal andrapidly dissociable complexes were measured as a Chelex-labilefraction. The metal concentrations were determined by ICP–MSand the Chelex-labile metal fractions were calculated as the differ-ence between the column influent and effluent concentrations.Method QA/QC consisted of running an inorganic (100% labile) cop-per solution at the beginning of each sample batch and a copper–nitrilotriacetic acid (NTA) solution that contained an average of14% Chelex-labile copper, at the end of each batch.

2.5. Colloidal metal analyses

Colloids have sizes in the range of approximately 0.001–1 lm(Wilkinson and Lead, 2007). It is well recognised that the filtrationof waters through a 0.45 lm filter (i.e. the standard filter size) doesnot completely separate dissolved and colloidal forms of metals(Gundersen and Steinnes, 2003; Wilkinson and Lead, 2007; vanLeeuwen and Buffle, 2009). In our study, after filtration through0.45 lm filters, water samples were then passed through an acid-washed 0.025 lm filter (Millipore), and the difference betweenthe 0.45 lm and 0.025 lm filterable metal concentrations wasattributed to colloidal metal forms. While <0.025 lm sized colloidsmay be present, it was not feasible to use smaller filter sizes. Dial-ysis was considered as an alternative technique, but was not useddue to concerns that the longer time periods required for dialysismay allow for undesired changes in dissolved and colloidal metalforms (Ure and Davidson, 1995; Guo et al., 2001). Method QA/QCconsisted of spiked Murray River water samples (<0.45 lm filtered,pH 7.8) containing 10, 50 and 100 lg L�1 of metals (Cd, Co, Cu, Ni,Pb, Zn); the samples were filtered and analysed to assess whetheradsorption of dissolved metals occurred onto the filter membranesand could influence the calculation of the colloidal metal fraction(Gardner and Hunt, 1981; Hedberg et al., 2011). The recoverieswere 88–108% of the nominal concentrations for the <0.45 lm fil-trates and 74–90% for the <0.025 lm filtrates for all metals except

lead. For lead, the recoveries were 44–62% for <0.45 lm filtratesand 36–37% for the <0.025 lm filtrates, and indicated that a con-siderable amount of lead was adsorbing to the cellulose acetatemembrane filters. For this reason, the accuracy of the dissolvedlead concentration was questionable and the lead results are notdiscussed further.

2.6. Dissolved metal concentrations and speciation following dilutionwith Murray River water

Dissolved metal concentrations were analysed in all six LMRIAdrainage waters and in 50%, 10% and 2% dilutions with Murray Riverwater. The dilutions were performed by rapidly adding the appropri-ate portions of the two waters (900 mL in total) into a 1-L bottle(LDPE), capping and then mixing by inverting several times and thenplacing on a purpose-built bottle roller at 30 rpm for the duration ofthe experiment. Subsamples were taken as required for analyses andthe bottles immediately returned to the roller. The dissolved oxygenconcentrations were maintained at >80% saturation during the tests.Changes in dissolved concentrations and speciation over a longertime period were also assessed for all LMRIA drainage water dilutedto 10% with Murray River water, and to 2% for the Mobilong water.The speciation in less dilute waters was not accessible due to theirlow pH preventing use of the Chelex method. The diluted sampleswere stored in a dark incubator at 18 ± 3 �C for the entirety of theexperiment to reflect summer water temperatures and low lightpenetration due to water turbidity. Three forms of metals weredetermined at time points over 8 d (0, 1 and 4 h, and 1, 3 and 8 d postmixing): Dissolved (the standard <0.45 lm filterable concentra-tion); Colloidal (the difference between <0.45 lm and <0.025 lm fil-terable concentrations); and Chelex-labile (free or weaklycomplexed and non-colloidal <0.45 lm filterable concentration).

3. Results and discussion

3.1. Drainage water composition

The drainage waters had pH values ranging from 2.9 to 5.7 andacidity ranging from 50 to 1190 mg L�1 CaCO3 (Table 1). The Tooraand Mobilong waters had high specific electrical conductivity(20–29 mS cm�1), which was consistent with the high chlorideand sulfate concentrations in these waters (Table 1). Dissolved oxy-gen concentrations were relatively low for the Long Flat and Wel-lington sites at the time of sampling the waters and will haveincreased before the filtration of the water occurred. Dissolved or-ganic carbon is an important metal-binding phase and had very highconcentrations in many of the drainage waters (Table 1). The drain-age waters with the lower pH usually had the higher dissolved me-tal concentrations (Tables 1 and 2; Fig. S2 of the Supplementaryinformation), however the soil properties and drainage processesat each site were also expected to have influenced the metal concen-trations. Consistent with a past study of metal mobilisation fromacidified soils in this region (Simpson et al., 2010), accompanyingthe high concentrations of dissolved Al (up to 55 mg L�1), Fe(205 mg L�1), and Mn (24.7 mg L�1) were high concentrations ofCo (up to 1.25 mg L�1), Ni (1.30 mg L�1) and Zn (1.10 mg L�1) (Ta-ble 2). Tables S1–S3 of Supplementary information section providethe water quality parameters and concentrations of less significantsubstances; Na, K, As, Ag, Cd, Cu and Cr. Hardness-adjusted WQGsfor Ag, Cd, Cr, Cu, Ni, Pb and Zn are provided in Table 2 for waterswith 30, 60 and 90 mg CaCO3 L�1 hardness (ANZECC/ARMCANZ,2000). Some of the drainage waters had salt concentrations exceed-ing those in seawater (e.g., Ca, Mg, and SO4), and the WQG valuesshould be considered as very low reliability when applied to theundiluted and highly acidic waters.

Table 1Major water quality data for sampled waters.

Water pH SEC DO Alkalinity Acidity DOC Chloride Sulfate FRP Ca MgmS cm�1 mg CaCO3 L�1 mg L�1 mg L�1 mg L�1

Murray River 7.8 0.43 5.2 95 <1 9 62 17 0.08 18 10Toora 4.8 20 5.7 <1 31 22 570 2170 <0.01 560 680Mobilong 2.8 29 6.3 <1 1070 37 9200 3710 0.03 790 1450Long Flat 5.3 3.8 2.0 4 100 14 650 720 <0.01 220 160Woods Point 3.1 8.5 7.0 <1 180 23 1370 1980 <0.01 450 460Wellington 3.4 4.2 2.7 <1 100 24 650 840 <0.01 200 200

SEC = Specific electrical conductance. DO = dissolved oxygen. FRP = Filterable reactive phosphate.Alkalinity and acidity can be converted to meq L�1 by dividing by 50.

176 S.L. Simpson et al. / Chemosphere 103 (2014) 172–180

Table 2Water pH and dissolved metal concentrations in undiluted waters and following immediate mixing of drainage and Murray River water.

Water Drainage water, %a pH Al Fe Mn Co Ni Znlg L�1

Murray River 0 7.8 <3 <3 <3 1.6 1.7 1.6

Toora 100 4.8 1100 170 5500 240 230 110Toora 50 6.9 51 31 2300 140 120 58Toora 10 7.7 22 5 480 25 28 9.0Toora 2 7.8 8 5 90 6 7.1 1.0

Mobilong 100 2.7 55000 205000 24700 1250 1300 1100Mobilong 50 3 2000 75000 9400 700 710 580Mobilong 10 5.4 870 105 2100 160 170 120Mobilong 2 7.3 28 16 420 31 34 12

Long Flat 100 4.9 30 90 3400 69 57 44Long Flat 50 6.4 8 15 1200 34 30 21Long Flat 10 7.5 6 5 240 7.0 6.8 3.0Long Flat 2 7.8 3 5 40 2.0 3.3 1.0

Woods Point 100 3.1 8900 18000 7400 260 370 270Woods Point 50 3.3 2000 630 3100 140 190 140Woods Point 10 7.2 47 16 630 25 40 17Woods Point 2 7.7 36 5 120 4 8.8 2.0

Wellington 100 3.3 4100 4600 5100 210 220 150Wellington 50 5.8 415 1150 2400 110 120 79Wellington 10 7.4 36 10 450 20 23 8.0Wellington 2 7.8 14 5 80 4 5.5 1.0

WQG (95%PC; TV �30 mg CaCO3 L�1)b 0.5/55 NV 1900 1.4 11 8WQG (hardness = 60 mg CaCO3 L�1)b 0.5/55 NV 1900 1.4 20 14WQG (hardness = 90 mg CaCO3 L�1)b 0.5/55 NV 1900 1.4 28 20

a Drainage water (%) represents dilution with Murray River water (i.e. 50 = 50% drainage water).b WQG (95%PC) = Trigger value (TV) for 95% species protection applicable to freshwaters of hardness 30 mg CaCO3 L�1, from ANZECC/ARMCANZ (2000). For Al, WQGs is for

waters with pH > 6.5. NV = no WQG exists. Hardness-adjusted WQGs for Ni and Zn applicable to fresh waters (ANZECC/ARMCANZ, 2000). Values are Bold when >WQG(hardness = 90).

3.2. Dissolved metal concentrations following mixing of drainage andMurray River water

Laboratory mixing experiments were used to assess changes inwater pH and dissolved (<0.45-lm filterable) metal concentrationsupon immediate (within 6 min of mixing) dilution with MurrayRiver water. Upon dilution of the pH 2–5 acid drainage water to10% with river water (10% drainage water: 90% river water), thepH increased to between 7.2 and 7.7 (approaching the river waterpH of 7.8) at most sites (Table 2). The exception to this was themost acidic Mobilong site where the 10% dilution increased to justpH 5.4 (Table 2). At dilutions to 2% drainage water (98% MurrayRiver water), the Mobilong water pH increased to 7.3 and to pH7.7–7.8 for the other drainage waters. This is consistent with fieldresults that show acid drainage water neutralisation is achieved ina localised mixing zone in the Murray River (EPA, 2013).

The pH increases occurring upon dilution of the drainage waterswith Murray River water resulted in large decreases in <0.45 lmaluminium and iron concentrations that were far greater than whatcould be attributed to dilution alone (Table 2; Fig. S3 and Table S4 of

the Supplementary information for data for other metals). This isconsistent with past observations and modelling that these ele-ments precipitate rapidly (to form hydrous aluminium and iron oxi-des, respectively) when acidic waters are neutralised by mixingwith alkaline waters (Balistrieri et al., 2007), although mixing con-ditions and potential formation of metastable complexes (Bertsch,1987) will also influence the observed aggregation rate. A small de-gree of precipitation was also evident for manganese, but the de-creases in concentration of the other elements closely matchedthose expected based on conservative dilution (Fig. S3). This is con-sistent with previous research that manganese typically persists asMn(II) except in more alkaline (pH > 8.5) and oxic environmentswhere the oxidation kinetics to form hydrous manganese oxidesare more favourable (Davies and Morgan, 1989). ‘Conservativebehaviour’ was observed after initial mixing for Co, Ni and Zn(Fig. S3). Field studies of the mixing of acid mine drainage with riverwater have found similar results for these metals, with non-conservative behaviour (due to adsorption to metal oxide phases)generally becoming more significant at greater dilutions of acidicdrainage water with ambient water (Balistrieri et al., 2007; Schemel

S.L. Simpson et al. / Chemosphere 103 (2014) 172–180 177

et al., 2007). In our study, the adsorption of these metals to thefreshly precipitated hydrous aluminium and iron oxides appearedto be minimal in the initial 6-min mixing period.

The changes in pH and <0.45 lm Al, Fe, Mn, Co, Ni and Zn con-centrations were investigated over a longer time period (196 h) fol-lowing mixing of acidic drainage and Murray River water (Fig. 2(Al, Fe and Mn) and Fig. 3 (Co, Ni, Zn) for 2% Mobilong, 10% Tooraand Woods Point; Figs. S4–S6 of the Supplementary informationfor 10% Mobilong, Long Flat and Wellington). The concentrationsof all <0.45 lm metals decreased with time following mixing ofthe two waters (Figs. 2 and 3). These decreases were not due topH which was quite stable over the 196 h period (varying by upto 0.2 pH units, changing from 7.68 to 7.60 for 10% Toora, 5.42 to4.94 for 10% Mobilong, 7.26 to 7.43 for 2% Mobilong, 7.45 to 7.57for 10% Long Flat, 7.23 to 7.39 for 10% Woods Point, and 7.41 to7.55 for 10% Wellington). The trends in the removal of aluminiumand iron over time were similar to that expected during colloidaggregation (initial high removal rates decreasing over time) (Ives,1978).

3.3. Metal speciation/forms

At each time point following mixing of drainage and MurrayRiver water, the total dissolved metal concentrations (i.e.<0.45 lm filterable concentration) was analysed in terms of frac-tions present in Chelex-labile forms, 0.025–0.45 lm colloidalforms, and dissolved organic complexes and <0.025 lm colloidalforms (combined) (Table 3). For all waters, Co, Mn and Zn werepredominantly present in Chelex-labile forms (95–100% of<0.45 lm fraction, Table 3), i.e. present as free metal ions or weakinorganic complexes that can be considered as bioavailable. In

Fig. 2. Changes in <0.45 lm, <0.025 lm and Chelex-labile Al, Fe and Mn concentrations10% Toora, and 10% Wood Point.

contrast, Al, Fe and Ni had a substantial fraction of the total<0.45 lm fraction (typically 60–80%) present in non Chelex-labileforms that were classified as either dissolved organic complexesor colloids (Table 3). The measurements indicated that there waslittle colloidal nickel present in the size range 0.025–0.45 lm,and the non Chelex-labile forms of nickel were either associatedwith colloids <0.025 lm or as organic complexes. The resultsmay be considered ambiguous as to whether the non Chelex-labilealuminium or iron was present as a mixture of organic complexesand colloidal forms, or as both larger and very small colloids. Thedifferences between aluminium or iron fractions at different sitesdid not demonstrate a clear relationship with initial water pH ormetal concentrations.

Previous studies investigating metal forms following mixing ofacid mine-drainage waters with ambient waters observed anapparent higher degree of colloid formation, e.g. 75% for alumin-ium (Gundersen and Steinnes, 2003; Schemel et al., 2007) and�40% for iron (Schemel et al., 2007). Due to the much higherDOC (10–40 mg L�1 compared to 0.2–2 mg L�1 in the acid mine-drainage studies), a higher amount of organic complexation wouldbe expected that may reduce the amount of colloid formation.However, Bigham et al. (1996) found that very small colloids (aver-age particle width 0.030 lm) were formed by precipitation in acidmine drainage. The potential influence of <0.025 lm colloids in ourresults is discussed further below.

The changes in metal form over 196 h are shown in Table 4(averaged for all the sites). Although there was considerable vari-ability between the portions of Chelex-labile Al, Fe and Ni in thedifferent waters, when the mean was calculated for each timeusing the combined data set for all the waters, the trend was quiteclear and consistent (Table 4, note that the 10% Mobilong dilution

following dilution of acid drainage waters with Murray River water: 2% Mobilong;

Fig. 3. Changes in < 0.45 lm and Chelex-labile Co, Ni, and Zn concentrations following dilutionof acid drainage waters with Murray River water: 2% Mobilong; 10% Toora, and10% Wood Point.

Table 3Fractions of the total dissolved metals in labile, colloidal or organically complexed forms.

Drainage water Forms of metals, % of <0.45 lm dissolved concentration

Chelex-labile (mostly free metal ions or weak inorganiccomplexes)

Colloidal (inorganic and organiccolloids)

Organically complexed and <0.025 lmcolloidal forms

Al Co Fe Mn Ni Zn Al Fe Ni Al Fe Ni

Toora 10% Mean 21 85 13 95 61 93 8 <2 3 71 98 36SD 14 18 7 3 11 16 11 32 3 6 29 12

Mobilong 2% Mean 38 100 39 98 62 95 25 15 3 41 46 35SD 16 0 8 2 11 10 28 18 3 26 19 12

Mobilong 10% Mean 69 96 61 95 99 98 2 <2 6 29 38 <2SD 7 3 7 2 3 1 9 16 3 7 14 3

Long Flat 10% Mean 26 70 42 99 38 87 57 40 4 25 18 59SD 13 10 11 2 7 15 37 28 2 45 26 8

Woods Point 10% Mean 39 96 20 98 68 100 28 11 6 34 69 26SD 6 3 6 2 9 0 26 20 4 31 18 11

Wellington 10% Mean 41 83 39 96 48 89 60 41 6 <2 20 47SD 6 16 15 4 8 17 15 20 14 16 17 12

All data, excluding10% Mobilong

Mean 33 87 31 97 55 93 39 19 4 28 50 43SD 13 15 15 3 14 13 30 30 7 34 38 13

Percentage of dissolved metals in Chelex-labile forms = [Chelex-labile]/[<0.45 lm], %.Percentage of dissolved metals in colloidal forms = ([<0.45 lm]–[<0.025 lm])/[<0.45 lm], %.Percentage of dissolved metals in organically complexed and <0.025 lm colloidal forms = 100 – %-Chelex-labile – %-Colloidal, %.Calculations were mean ± standard deviation of the 6 time points and did not include data with dissolved concentrations <25 lg L�1 Al, Fe, Mn or <5 lg L�1 Co, Ni, Zn.

178 S.L. Simpson et al. / Chemosphere 103 (2014) 172–180

was excluded from the mean calculations due to its much lower pHthan the other sites). Only for nickel was there a significant changein dissolved metal forms through time following mixing of theacidic drainage water with the Murray River water. The degree of

nickel association with either fine colloids (e.g. adsorption to<0.025 lm iron and manganese oxyhydroxide phases) or complex-ation by organic matter clearly increased with time, from 33 ± 12%(mean ± standard deviation) after 6 min to 54 ± 7% after 196 h. This

Table 4Changes in complexed or colloidal forms through time following mixing.

Time, h Metals in complexed or colloidal forms, % of total dissolved

Al Co Fe Mn Ni Zn

0.1 Mean 69 9 59 3 33 1SD 16 10 17 2 12 3

1 Mean 71 9 72 4 40 <1SD 12 9 13 2 16

4 Mean 68 9 73 5 42 <1SD 14 10 16 2 14

24 Mean 63 6 76 1 45 <1SD 12 5 14 1 12

72 Mean 65 6 72 <1 48 <1SD 17 10 16 11

196 Mean 63 <1 66 <1 54 <1SD 0 17 7

All Mean 67 8 69 3 43 <1SD 13 8 15 2 13

% Complexed or Colloidal = ([<0.45 lm] – [Chelex-labile])/[<0.45 lm], %.Data did not include 10% Mobilong or those with <25 lg L�1 Al, Fe, Mn or <5 lg L�1 Co, Ni, Zn.

S.L. Simpson et al. / Chemosphere 103 (2014) 172–180 179

may be attributed to aggregation of <0.25 lm colloidal nickelforms to form larger colloids. The total dissolved nickelconcentrations were relatively high and nickel is not consideredto be strongly complexed by organic ligands. For Al, Co, Fe, andZn little change occurred to the dissolved metal forms with time,with the concentrations of Chelex-labile and colloidal metalsdecreasing proportionally as the dissolved (<0.45 lm filterable)concentration decreased. It is suggested that this indicates<0.025 lm colloids may not be important and instead thatdissolved metal–organic complexes dominate this fraction. If finecolloids were predominant, they would have been expected toaggregate to form larger colloids then settle from solution overtime as discussed above. This would have reduced the concentra-tion of these metals in the <0.025 lm fraction, but this was notwhat was observed.

3.4. Water quality guidelines and potential toxicity

The dissolved concentrations of many metals in the drainagewaters greatly exceeded WQGs for 95% species protection(95%PC) (ANZECC/ARMCANZ, 2000) (Tables 2 and S4). Based onthe ratio of the metal concentration to WQG, and considerationof the Chelex-labile metal forms (Table 3), the metals of greatestconcern to the Lower Murray River ecosystem were determinedto be Al, Co, Ni and Zn. While the high manganese concentrationswere below WQGs for ecosystem protection (Table 2), they mayrepresent a potential risk to drinking water sources as followingdilution concentrations were often still well above guideline values(100 and 500 lg L�1 for aesthetic and health protection of drinkingwater respectively, NHMRC/NRMMC, 2011). All the metals exceed-ing WQGs have been previously identified as a particular concernin relation to release following oxidation of acid sulfate soils(Simpson et al., 2010; Mosley et al., 2013).

Following dilution and mixing with the ambient Murray Riverwaters, the dissolved metal concentrations dropped considerablywithin 6 min (Table 3). For the 2% mixture of the Mobilong waterand 10% mixtures of the other waters (with River Murray water),the WQGs were exceeded for cobalt in all waters. However, whilethe dissolved concentration of nickel also exceeded the WQG formost waters, the portion of Chelex-labile, and potentially bioavail-able nickel was below the WQG after 72 h (Table 4). The WQG forcobalt is very low at 1.4 lg L�1 and the cobalt concentration re-mained above this value for up to 8 d following mixing for manyof the mixed waters. It is noted that the WQG for cobalt is

considered to be of low reliability due to the very limited data onthe chronic toxicity of cobalt to freshwater biota (ANZECC/ARM-CANZ, 2000).

After 72 h, the dissolved concentrations of Mn, Ni and Zn hadfurther decreased, by approximately 4-fold and were within a fac-tor of two above or were below the WQG concentrations (Figs. 2and 3, and Figs. S4–S6 of the Supplementary information). Differ-ences in DO concentrations were determined to be a minor factorin determining the likelihood of exceeding the WQGs.

Regardless of the form, the Chelex-labile aluminium concentra-tions were less than the WQGs for waters with pH > 7. As a conse-quence, at pH > 7, the dissolved aluminium is expected torepresent a minimal hazard to the aquatic organisms. Greater pHvalues than this were observed with 50:1 dilutions of the aciddrainage waters with Murray River water (2%) and at present thisis achieved in the localised mixing zone within approximately20 m of the discharge (EPA, 2013). The potential toxicity of the me-tal-rich precipitates formed upon dilution of the acidic water re-quires further research (Gensemer and Playle, 1999; Cain et al.,2013).

When all of the dissolved metals data were considered together(dilution and kinetics experiments), a majority of the metals werehighly correlated (Table S5 of the Supplementary information) andall concentrations decreased steadily with increasing pH (Fig. S7 ofthe Supplementary information). For copper and lead, the dis-solved concentrations were generally low and the concentrationsclose to the limit of reporting will have contributed to the weakercorrelations. The strong correlations indicated that the factors thatgovern the removal of metals from solution during mixing of drain-age water with Murray River water were very similar for eachmetal.

The study indicates that for dilutions of the acidic drainagewaters to less than 1% with Murray River water, the dissolved met-als should not pose a risk of adverse effects to ecosystem health,with the possible exception of cobalt. The principal factors influenc-ing the dissolved metal concentrations were dilution factor, pH, andthe time following mixing, and these will be important to considerwhen developing management plans for the discharges and poten-tial future river conditions. Liming of the drainage channels to re-duce the acidity and precipitate metals prior to discharge may bea useful management option. An outstanding issue in relation tothe discharges is the response of organism to intermittent expo-sures to contaminants from the fluctuating or pulsed discharge ofthe drainage waters (Nordstrom, 2011). However, it is expected

180 S.L. Simpson et al. / Chemosphere 103 (2014) 172–180

that the time-averaged concentration of metals provides adequateprotection when comparing to WQGs (Angel et al., 2010).

The metal-rich precipitates that form in situ upon neutralisa-tion of the drainage waters in-river, or that may be dischargedfrom drains, represent an additional risk to some organisms. Anumber of studies have highlighted the importance of consideringthe dietary exposure of metals to benthic organisms, includingmetals adsorbed to colloidal hydrous iron oxides in river systems(Cain et al., 2013) and metals that accumulate within sediments(Campana et al., 2013). Further research is required to assess riskposed to benthic organisms by the metal-rich precipitates that re-main suspended or deposit within the river system and in situmeasurements within the river system to validate the laboratoryfindings (Liber et al., 2007).

Acknowledgements

This research was funded by the CSIRO Water for a HealthyCountry Flagship, the Environment Protection Authority of SouthAustralia (EPA) and the Murray Darling Basin Authority (MDBA).David Palmer from the EPA is thanked for assistance in the fieldwork and Josh King for assistance with metal analyses. The authorswould like to thank Graeme Batley and Lisa Golding for editorialcomments on this manuscript.

Appendix A. Supplementary material

Supplementary data associated with this article can be found, inthe online version, at http://dx.doi.org/10.1016/j.chemosphere.2013.11.059.

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