Applied Geochemistry 51 (2014) 33–43

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Applied Geochemistry

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Remediation of metal-contaminated soil in polar environments:Phosphate fixation at Casey Station, East Antarctica

http://dx.doi.org/10.1016/j.apgeochem.2014.08.0110883-2927/� 2014 Elsevier Ltd. All rights reserved.

⇑ Corresponding author. Tel.: +61 403 421 601.E-mail addresses: erla.hafsteinsdottir@gmail.com (E.G. Hafsteinsdóttir), kirstie.

fryirs@mq.edu.au (K.A. Fryirs), scott.stark@aad.gov.au (S.C. Stark), damian.gore@mq.edu.au (D.B. Gore).

Erla G. Hafsteinsdóttir a,⇑, Kirstie A. Fryirs a, Scott C. Stark b, Damian B. Gore a

a Department of Environment and Geography, Macquarie University, North Ryde, NSW 2109, Australiab Environmental Protection and Change, Australian Antarctic Division, Department of Sustainability, Environment, Water, Population and Communities, Kingston, Tasmania7050, Australia

a r t i c l e i n f o

Article history:Available online 28 September 2014Editorial handling by M. Kersten

a b s t r a c t

Remediation of metal-contaminated soils by phosphate fixation is successful in temperate environments,whereas its efficacy in cold and freezing environments is understudied. Phosphate fixation is a low-costtechnique and is potentially very useful in these remote environments where the logistics of remediationare difficult and expensive. Here we describe a field study at Casey Station, East Antarctica, where phos-phate (triple superphosphate and phosphate rock) and a buffer, Emag (magnesium carbonate and mag-nesium oxide), were introduced to contaminated soil from nearby Thala Valley landfill. A pilot scaleexperiment was set up, sampled and monitored from December 2008 to February 2010. Relative to levelsin the untreated landfill material, concentrations of Cd, Cu, Co, Fe, Mn, Pb and Zn in leachates weredecreased by phosphate addition (fixation most effective for Mn and Zn), whereas the techniqueincreased concentrations of As, Cr and Ni. The most successful fixation was found for the 3:2 ratio of triplesuperphosphate and Emag, and the least effective fixation occurred with the 2:1:1 ratio of triple super-phosphate, Emag and phosphate rock. Although there was an undesirable initial flush of metals from thecontaminated soil in the 24–48 h after treatment addition, concentrations in leachate were reduced andstabilised in the second summer. During a full-scale field implementation, complementary techniqueswould be required to contain and treat contaminated runoff until leachates have reduced to acceptableconcentrations.

� 2014 Elsevier Ltd. All rights reserved.

1. Introduction

Antarctica, the coldest, driest and most remote continent, hasexperienced a major increase in human activity over the last50 years following the International Geophysical Year (IGY) in1958–1959, when 55 research stations commenced operation(Bargagli, 2008). Early waste management consisted mainly ofincineration, dumping of domestic, construction, and scientificwaste in landfills near the stations, or disposing waste onto thesea ice where it sank into the sea or dispersed upon sea ice break-age (Deprez et al., 1999; Snape et al., 2001b; Stark et al., 2006;Bargagli, 2008). Consequently, many contaminated sites are nowlocated around the continent, mostly on the coast, where freeze–thaw cycling of soil and meltwater runoff into the adjacent marineenvironment occurs (Deprez et al., 1999; Snape et al., 2001b; Starket al., 2003a; Fryirs et al., 2013, 2014). This is of great concern as

the terrestrial and marine ecosystems around the continent maybe especially vulnerable to contamination due to their uniqueadaptation to the harsh, isolated, Antarctic conditions (Bargagli,2008).

Metal contamination in landfills is a concern at Australian Ant-arctic stations (Deprez et al., 1999; Snape et al., 2001b; Stark et al.,2003b; Fryirs et al., 2013). Landfills contain complex mixtures ofmetals that can create environmental risk if allowed to leach off-site (Snape et al., 2001b; Stark et al., 2003a, 2005). One optionfor remediation is phosphate fixation of on-site materials. Thistechnique is relatively low-cost, and can be used readily in remoteenvironments where the logistics of removal, transport and reloca-tion of materials are difficult. Phosphate fixation transforms metalsinto insoluble or sparingly soluble forms (Bolan et al., 2003;Hettiarachchi et al., 2000; Porter et al., 2004; Scheckel and Ryan,2004), reducing the environmental risk of landfill leachates. Reme-diation of metal contaminants by phosphate is routine in temper-ate environments (Melamed et al., 2003; Wang et al., 2001; Yangand Mosby, 2006), but the method remains little tested in coldand freezing environments. Freezing and thawing of near-surfacelayers can potentially affect the stability of the fixation products

34 E.G. Hafsteinsdóttir et al. / Applied Geochemistry 51 (2014) 33–43

as freezing desiccates soil particles and increases solute concentra-tion in the thin liquid film that remains around particles (Blackwellet al., 2010). Laboratory trials of the technique have been con-ducted to stabilise metal contaminated landfill deposits in coldenvironments (Hafsteinsdóttir et al., 2011, 2013; White et al.,2012). Laboratory studies in simple model systems (i.e. not realsoils) demonstrated successful formation and stability of Cu-,Pb- and Zn-phosphate minerals during a simulation of 20 yearsof freeze–thaw cycles, which are experienced at 30 cm depth inthe soil at Casey Station, East Antarctica, demonstrating significantpromise for the technique in cold environments (Hafsteinsdóttiret al., 2011, 2013). However, laboratory experiments are simplifi-cations of the natural environment and field verification is critical.

Thala Valley landfill, adjacent to Brown Bay and <1 km fromCasey (Fig. 1), operated between 1965 and 1986 (Deprez et al.,1999). The landfill stored steel drums, discarded machinery, vehi-cle parts, batteries, pipes, mattress springs, rusted cans and toast-ers, food, photographic chemicals and hydrogen generation waste(Deprez et al., 1999; Snape et al., 2001b) in a sandy to gravellymatrix of weathered rock and glacial debris (Deprez et al., 1999;Thums et al., 2010). The waste included �11 wt% organic matter,consisting mainly of food scraps and wood (Northcott et al.,2003). The surrounding landscape consists of low-lying hills ofbedrock and gravel, with mosses, lichens, and algae as the onlyvegetation (Ling, 1996; Melick and Seppelt, 1997; Smith, 1990).The area experiences a mean annual snowfall around 195 mm/year(rainfall equivalent) (Melick and Seppelt, 1997), and mean summerand winter air temperatures of �3 and �12 �C, respectively(Australian Antarctic Program, 2011). The site is buried under snowand ice for most of the year and is underlain by permafrost. Duringsummer the land is exposed from the snow, allowing surface soilsthat overly the permafrost to thaw (Deprez et al., 1999). Commonlythis surface layer freezes at night and thaws during the day, gener-ating meltwater that flows through the landfill site and mobilisescontaminants.

Dissolution of metals and mobilisation of contaminated finesediment by meltwater flowing through the surface and subsurfaceof Thala Valley landfill has caused elevated concentrations of Cd,Cr, Cu, Fe, Mn, Ni, Pb, Sb, Sn, Zn in the adjacent bay waters and sed-iments (Deprez et al., 1999; Snape et al., 2001b; Stark et al.,2003a, 2005). As a result, the landfill was assessed as a site of

Fig. 1. Casey Station area, Windmill Islands, East Antarctica. Thala Valley landfill leacheestablished, mixing �3.6 tonnes of landfill material with phosphate, in an effort to mini

environmental concern, and assigned the highest priority statusfor remediation of Australian Antarctic contaminated sites(Deprez et al., 1999; Snape et al., 2001b). To ensure that Australiamet its environmental obligations under the Madrid Protocol of1991 (ATCPs, 1993), the goals of the research at Thala Valley wereto develop and implement cleanup techniques that have no greateradverse environmental impact than leaving the waste undisturbed(Stark et al., 2006).

The landfill was subjected to a staged dig-and-haul cleanupfrom 1995 to 2011 (Deprez et al., 1999; Snape et al., 2001a,2001b; Stark et al., 2006, 2008). In December 2008, an on-sitepilot scale experiment was established at Casey using stockpiledlandfill material. Some macro-waste objects were collected inFebruary 2010 for identification of contaminant sources(White et al., 2012). Total and leachable metal concentrations insoil indicated Pb as the main contaminant of concern, while Cd,Cu, Ni and Zn posed a lower environmental risk (Thums et al.,2010).

For the on-site experiment, phosphate (phosphate rock and tri-ple superphosphate) was introduced to 3.6 t of the stockpile mate-rial in December 2008, and the experiment was sampled andmonitored through to February 2010, providing a 14-month recordof treatment effectiveness. This allowed the materials to undergonatural weathering conditions and freeze–thaw cycling. The aimsof this study were to determine how efficient on-site phosphatefixation is at treating contaminated landfill material, and whatratio of reagents is required for most effective treatment in a coldenvironment. If successful, this technology could be used to reme-diate metal contaminated sites in Antarctica and other cold andfreezing environments.

2. Materials and methods

2.1. Experiment design, set-up and instrumentation

During the 2008/09 field season, Thala Valley landfill materialwas collected from a stockpile on site. Samples were taken fromthe surface and internal parts of the stockpile, sieved at 50 mmand placed into a shipping container. Macro-waste objects <30 cmdiameter were retained in the sample fill (Fig. 1Sa (Supporting

d metals into adjacent Brown Bay. In December 2008 a pilot scale experiment wasmise the leaching of metals from the landfill waste.

E.G. Hafsteinsdóttir et al. / Applied Geochemistry 51 (2014) 33–43 35

Information), whereas larger pieces were discarded. Clean crushedrock from the Casey quarry was used as a background control.

Three shipping containers (3 � 2.4 � 1.2 m; C1–C3, Fig. 2), eachcontaining six, well-washed 205 L steel drums (h = 96 cm,r = 29 cm; D1–D6), were installed on-station (Fig. 1). The contain-ers prevented spillage of contaminants into the environment. Con-tainer 1 (C1: 6 drums) contained six fully instrumented drums.Containers 2 and 3 (C2 and C3: 12 drums) were used for replicatesampling and contained less instrumentation (Fig. 3).

Leachate samplers were constructed from 50 mm slotted PVCtubes attached to PVC electrical conduit that acted as an accesstube. Plastic tubing was inserted into the conduit and the slottedPVC tube that sat at the bottom of the drum. A 3-way valve tapwas attached to the tube from which leachate samples were col-lected using a plastic syringe.

Soil temperature and moisture were monitored using EC-TMsensors connected to Em50 data loggers (ECH2O�; DecagonDevices Inc., USA), and a PR2/6 profile probe connected to ahandheld DeltaLINK HH2 moisture meter (Delta-T Devices Ltd.,UK). Two EC-TM soil temperature (�C) and moisture (volumetricwater content (VWC); m3/m3) sensors were attached to electricalconduit at 30 and 60 cm above the bottom of the drums in con-tainer 1 (Fig. 3a), using an Em50 self-contained data recorderwith 5 sensor ports and one communications port taking readingsevery 2 h. ECH2O software configured the data logger andexported the data to MS Excel. The VWC range was 0–1 m3/m3 ± 3%, and temperature was �40 �C to +50 �C ± 1 �C.

A handheld DeltaLINK HH2 moisture meter connected with aPRC/d-HH2 cable to a PR2/6 profile probe (with sensors at 6 depths)was used as backup for datalogged EC-TM data. The probe’s fulloperating range was 0.0–1.0 m3/m3 at �20 to +70 �C; accuracy forthe range 0.0–0.4 m3/m3 was ±0.06 m3/m3 at 0 to +40 �C. Threeaccess tubes were installed in the centre (M1), middle (M2) andedge (M3) of each drum in container 1 (Fig. 3a) and in the centre(M1) of containers 2 and 3 (Fig. 3b).

The contaminated stockpile material was treated with triplesuperphosphate fertiliser (TSP; Ca(H2PO4)2�H2O), phosphate rock

Plan view

D5 D6

Heat D3 D4

D1 D2

tracebox

C1

C3 C2

ND1 Untreated landfill – control

Quarry – background controlD2

D3 Landfill material mixed with TSP:emag – 2:1 (total 3% w/w)

D4 Landfill material mixed withTSP:emag:PR – 2:1:1 (total 4% w/w)

D5 Landfill material mixed withTSP:emag – 3:2 (total 5% w/w)

D6 Landfill material mixed withTSP:emag:PR – 3:2:2 (total 7% w/w)

Fig. 2. Experimental design, composition of the treatments and instrumentation in each(Ca(H2PO4)2�H2O), PR = Phosphate Rock (Ca5(PO4)3F), Emag = mixture of MgO and MgCO

(fluorapatite, Ca5(PO4)3F) and Emag (caustic calcined magnesia, amixture of MgO and MgCO3) (Fig. 2). Triple superphosphate acidi-fies the material being fixed (Hettiarachchi et al., 2001), enhancingdissolution of contaminant-bearing phases and enabling metal andphosphate reactions. Phosphate rock is less soluble (Knox et al.,2006), providing a slower release phosphate source than TSP. Emagis a pH buffer. These phosphate sources do not invoke Antarcticquarantine restrictions, unlike biogenic phosphate sources suchas fishbone.

Two controls and four treatments were used in each container.D1 was untreated landfill material (control) and D2 was quarrymaterial (background control). The remaining drums (D3–D6) hadlandfill material mixed with treatment materials at different ratios(Fig. 2). Selection of the treatment ratios was based on analyticaldata for aqueous extracts of Thala Valley landfill samples preparedusing USEPA Method 1311 – the Toxicity Characteristic LeachingProcedure (USEPA, 1992; Stark et al., 2008), and unpublished infor-mation on treatment ratios used by industry partner Veolia Environ-mental Services. The amount of treatment reagent was based on thedry weight of the landfill material (�300 kg lightly packed into eachdrum, with estimated 1.65 kg dm�3 density and 20% porosity).Moisture content of the landfill material, measured by oven dryingat 105 �C for 24 h, averaged 7.3% (range 6.9–7.8%), consistent withprevious analyses (Northcott et al., 2003).

Treatment reagents for each drum had been pre-weighed intoplastic bags, ready for mixing with landfill material. A concretemixer was used to homogenise the landfill material and the treat-ments. This method is standard industry practice, with concretetruck mixers used when large volumes are treated. After mixing,the material was poured into the drums (Fig. 1Sb (SupportingInformation)). At 11:30 on 22 December 2008, 60 L of clean snow-melt water was added to each drum with landfill material (D1,D3-6) and 40 L to the drum with background control (quarry;D2) to start the experiment and simulate the summer melt. Thequarry material was finer and had a lower water holding capacity.No additional melt water was added in 2009/10, but melt fromsnow fall contributed free water to the drums.

C1: •Access tubes for PR2/6 soil moisture profile probe (x3 in each drum)

D5 D6(x3•Leachate tubes with heat trace wrapped around (x1 in each drum)•EC-TM sensors (30 & 60

D1

D3 D4

D2

cm above bottom)

C2: •Access tubes for PR2/6 soil moisture profile probesoil(x1 in each drum)•Leachate tubes with heat trace wrapped around (x1 in each drum)

D4

D6

D3

D5C3: •Access tubes for PR2/6 soil moisture profile probe (x1 in each drum)

D1 D2

( )•Leachate tubes with heat trace wrapped around (x1 in each drum)

Casing for data loggers

drum (D1–D6), in each of the three containers (C1–C3). TSP = Triple Superphosphate3, w/w = weight/weight (dry).

For collecting leachate samples

Access tubes for PR2/6 soil moisture profile probe

Temperature and moisture sensors (EC-TM)

Ele

ctric

al c

ondu

it w

ith h

eat t

race

Slotted PVC with heat trace

30 cm

60 cm

Ele

ctric

al c

ondu

it

M1 M2 M3

M2 M3

LeachateEC-TM

NPlan view

Cross section view

For collecting leachate samples

Access tube for PR2/6 soil moisture profile probe

Ele

ctric

al c

ondu

it w

ith h

eat t

race

Slotted PVC with heat trace

M1

M1

Leachate

N

Cross section view

Plan view

(a) (b)

M1

Fig. 3. Instrumentation for drums in (a) container 1, and (b) containers 2 and 3. Drums in container 1 contained all the instrumentation and data loggers; drums in containers2 and 3 had less instrumentation.

36 E.G. Hafsteinsdóttir et al. / Applied Geochemistry 51 (2014) 33–43

2.1.1. Leachate sample collection and analytical methodsLeachate samples were collected over 1 month during the 2008/

09 and 2009/10 summers (Table 1). Duplicate samples were col-lected from each drum in container C1 to assess the effect of sam-pling precision on analytical uncertainty for leachates. To evaluateuncertainty associated with the experimental setup, including het-erogeneity in composition of the landfill or control material andapplication of the treatment reagents, data for leachates fromdrums in containers C2 and C3 were combined with data fromC1 to calculate mean (n = 3) values for each control or treatmentat the different sampling times (see Supporting Information). Plas-tic syringes were used to withdraw leachate samples through plas-tic tubing from the bottom of each of the 18 drums. Samples werefiltered immediately with 47 mm diameter cellulose acetate0.45 lm membrane syringe filters (Sartorius), and placed into50 mL polypropylene sample vials (Sarstedt). Samples were acidi-fied to 1% v/v (pH < 2) with analytical grade concentrated HNO3

Table 1Times for leachate sample collection. The experiment started at 11:30, 22ndDecember 2008, when meltwater was added to each drum. T = 0 samples werecollected when sufficient water had moved through to the bottom of the drum.Sample collection was frequent during the first 24 h of the experiment and thenspaced over intervals of P2 d.

Sample collection

Date Time

2008/09T = 0 (3 h) 22 December 2008 14:15––15:15T = 1 (7 h) 22 December 2008 18:15–19:00T = 2 (12 h) 22 December 2008 23:00–00:25T = 3 (24 h) 23 December 2008 11:50–12:30T = 4 (48 h) 24 December 2008 11:40–12:20T = 5 (96 h) 26 December 2008 12:15–13:15T = 6 (144 h) 28 December 2008 11:40–12:25T = 7 (192 h) 30 December 2008 11:30–12:30T = 8 (240 h) 01 January 2009 11:40–13:05T = 9 (432 h) 09 January 2009 15:00–16:15T = 10 (768 h) 23 January 2009 14:40–15:30

2009/10T = 11 07 January 2010 11:30–13:00T = 12 09 January 2010 8:30–10:00T = 13 11 January 2010 8:30–9:30T = 14 13 January 2010 8:30–9:20T = 15 25 January 2010 8:30–9:50T = 16 27 January 2010 8:30–9:15T = 17 29 January 2010 9:00–10:10T = 18 01 February 2010 8:30–9:20T = 19 03 February 2010 8:40–9:45T = 20 05 February 2010 8:45–9:45

(BDH Aristar) and stored at +4 �C. At each sampling time, an addi-tional sample from each drum was collected for measurement ofoxidation-reduction potential (Eh), electrical conductivity (EC)and pH. These measurements were carried out using electrodesin the Casey laboratory within 6 h of sample collection.

Preserved leachate samples were returned to Australia for anal-ysis of As, Ca, Cd, Co, Cr, Cu, Fe, Mg, Mn, Ni, P, Pb, Sb, Se, Sn, and Znby Inductively Coupled Plasma Optical Emission Spectroscopy(ICP-OES), using a Varian 720-ES spectrometer at the AustralianAntarctic Division, within 3 months of collection. The emissionintensity of standards and sample solutions was measured follow-ing 1:1 in-line mixing with a solution of 0.75% (w/v) CsCl (matrixmodifier) in 10% (v/v) HNO3 containing 5 mg/L yttrium as internalstandard. The method detection limit (MDL), evaluated from repli-cate measurements of 20 lg/L standards, was employed as thereporting limit for leachate data from both seasons (Table 1S)(see Supporting Information for further information on analyticalquality control).

3. Results

3.1. Soil moisture and temperature

The EC-TM temperature sensors recorded a similar trend foruntreated (D1-2) and treated (D3-6) material. For the length ofrecording, the average minimum and maximum temperaturesrecorded at the top of the drum were �21 ± 1 �C and +18 ± 1 �Crespectively, and at the bottom of the drum were �25 ± 1 �C and+12 ± 1 �C respectively. All systems (D1-6) showed highest mois-ture content during the summer, with moisture content nearingzero during the winter months (Fig. 4), consistent with conversionof most liquid water to ice. This indicates that phosphate fixationof the metal contaminated material should occur for at least4 months of the year until the desiccating effects of freezing duringthe winter months slow or cease the reactions.

Soil moisture was recorded at 140 mm and 540 mm depth(Fig. 5). In the control drums (D1-2), moisture remained stablebetween seasons throughout the drum and an initial increase inmoisture content was not observed during the first hours of theexperiment in 2008/09. During the 2008/09 summer, moisture con-tent at both the top and bottom of the drum varied more in the trea-ted landfill (D3-6) compared to the controls. During the 2009/10summer, the moisture content remained stable. Moisture contentin the treated materials (D3-6) decreased from the time the firstleachate sample was collected, particularly towards the top of thedrum (140 mm depth) (Fig. 5).

-0.1

0.0

0.1

0.2

0.3

0.4

0.5

0.6

18-Dec-08 07-May-09 25-Sep-09 12-Feb-10

VW

C (

m3 /

m3 )

Date

D1: 30 cm D1: 60 cm D2: 30 cm D2: 60 cm D3: 30 cm D3: 60 cm D4: 30 cm D4: 60 cm D5: 30 cm D5: 60 cm D6: 30 cm D6: 60 cm

Fig. 4. Volumetric water content in all drums at 30 and 60 cm above the bottom,from 20 December 2008 to 12 February 2010. All drums show similar trends withthe greatest water content over the summer, indicating that the treatment shouldwork for at least 4 months until the freezing winter temperatures start. Wintervalues for two probes are negative, reflecting a calibration error of �0.03 m3/m3.This figure emphasises the winter–summer trend, with higher moisture contentduring the summer months compared with the winter months.

E.G. Hafsteinsdóttir et al. / Applied Geochemistry 51 (2014) 33–43 37

3.2. Electrical conductivity, pH and Eh

In 2008/09, the pH of leachates ranged from 5.09 to 7.73 (Fig. 6).The highest pH was recorded for the background (7.42–7.73) anduntreated landfill (6.95–7.53) controls, whereas the lowest pHwas recorded for treatment D4 (5.09–5.70). Of the four treatments,the landfill material with the higher ratio treatments (D5 and D6)showed a larger range in pH (5.88–6.54 and 5.57–6.19, respec-tively). Over the first 48 h of the experiment, pH decreased in all

Fig. 5. Soil moisture during sample collections in 2008/09 (left) and 2009/10 (right). Moisperiod (800 h) in both graphs.

treated materials (D3-6) before increasing again. The smallest ini-tial decrease in pH was recorded for the landfill material mixedwith the higher ratio treatments (D5 and D6). After �1 month,the pH for treatment D5 was approximating the pH of theuntreated landfill and by 2009/10, treatment D5 had the highestpH value. The remaining treatments (D1-4 and D6) had all stabi-lised to within a similar range. The pH values remained constantthroughout 2009/10 (Fig. 6).

In 2008/09, following addition of well-oxygenated melt waterto the experiment, Eh values showed an initial increase in all sys-tems (D1-6), reaching maxima after 24 h (+108 to +194 mV). Fol-lowing the maxima, Eh for all systems decreased. In 2008/09,average Eh values for the treated materials (D3-6) were all positive(+41 to +194 mV). The controls (D1-2) reached negative Eh valuesafter �432 h (18 d). In 2009/10, a similar trend in Eh was observedfor all systems (D1-6), with values mainly remaining negative(Fig. 6), and a strong sulphide odour from the samples consistentwith reducing conditions.

Electrical conductivity for the background and untreated land-fill controls remained relatively constant throughout 2008/09 and2009/10. However, leachates from the treated materials (D3-6)had an initial increase in EC, reaching maxima after 96–192 h. Aslight decrease then occurred and EC remained constant through-out 2009/10. Almost identical EC values were observed for theuntreated landfill control and the landfill material mixed withthe D5 treatment (Fig. 6).

3.3. Leachable metals

Currently there are no generalised environmental guidelines setto assess the potential risk of contaminants in Antarctic ecosys-tems. Therefore, it is necessary to compare contaminant concentra-tion data with guidelines derived for other (typically temperate)

ture was measured at 140 mm and 540 mm depths. X-axes represent the same time

Fig. 6. Average pH, Eh and EC for seasons 2008/09 and 2009/10. X-axes represent the same time period (800 h) in both graphs. Note different y-axis scales for the 2008/09 and2009/10 season Eh charts.

38 E.G. Hafsteinsdóttir et al. / Applied Geochemistry 51 (2014) 33–43

environments (e.g. the Australian and New Zealand Guidelines forFresh and Marine Water Quality (ANZECC/ARMCANZ, 2000)), and/or site-specific background values. In this investigation, back-ground levels determined using the quarry soil exceededANZECC/ARMCANZ (2000) guidelines for many of the analytes(As, Cd, Cr, Cu, Ni, Pb and Se) and so direct comparison of leachateswith these guidelines is of limited value (Table 1S (SupportingInformation)). Rather, comparison with site-specific values, i.e.quarry background control (D2) and untreated landfill control(D1) data collected during the 2008/09 and 2009/10 seasons, ismore useful for evaluating the results presented here.

Samples were collected for the analysis of leachable metals over1 month during the 2008/09 and 2009/10 summers (Fig. 2S(Supporting Information)). Concentrations of leachable metals inthe background (quarry material, D2) and untreated landfill (D1)controls were mostly < 100 lg/L (Table 1S (Supporting Informa-tion)), with Cd, Sb, Se, and Sn, close to or below the reporting lim-its. The background control typically produced the lowestconcentrations of leachable metals, as expected. The untreatedlandfill control had, in most cases, lower and less variable leach-able metal concentrations than the treated material during the2008/09 season. It maintained these through 2009/10, providing

a benchmark against which to test the effectiveness of differenttreatments at reducing metal leachate concentrations.

In marked contrast to the controls, a significant increase or‘pulse’ in concentration was noted for almost all metals in the first24–48 h after addition of water to the treatments, with some met-als reaching their maximum concentration after 8 d of reactiontime (subject to treatment type). Depending on the metal, theseconcentrations did not recede to initial levels until at least 6 d aftertreatment, or even after �1 month for some metals (Co, Cu, Sb).

Owing to logistical and practical constraints, controls constitut-ing the background quarry material (D1) or a substitute (e.g. acid-washed sand) mixed with the treatment reagents were notincluded in the experiment. Therefore we cannot quantify the con-tribution of trace elements in the treatment materials to leachateconcentrations, and in particular, to those measured during thecontaminant pulse. Certainly the treatment reagents used (includ-ing the feed water added to the drums) contained contaminants attrace concentrations (see Supporting Information), and leachingfrom these types of reagent, albeit at low levels, is well established(Knox et al., 2006). However, while we cannot rule out some con-tribution of contaminants from the treatment reagents, there is noconsistent pattern in the contaminant profiles (Fig. 2S (Supporting

E.G. Hafsteinsdóttir et al. / Applied Geochemistry 51 (2014) 33–43 39

Information)) indicating that this was the major causative factor ofthe concentration pulse. If contamination from treatment chemi-cals were responsible, then (ignoring pH effects arising from differ-ing ratios of Emag) one would predict concentration maxima foreach element to be correlated with the wt% of treatment chemicals(i.e. D6 > D5 > D4 > D3). This was not observed. Also, we cannotattribute contamination from the feed water as the cause becausethis water was used for all the drums, including the controls, forwhich a contaminant pulse was not discernible. Furthermore, themaximum concentrations of contaminants measured were notatypical of aqueous extracts of material sourced from the ThalaValley tip (Snape et al., 2001b; Stark et al., 2008 and unpublished),including the stockpile from which experimental material wasgathered (Thums et al., 2010) (Table 1S (Supporting Information)).

With the exception of Fe and Sn, all treated materials producedleachates with concentrations of metals that remained above thebackground control (D2) during the first month of reaction (Snwas typically close to the MDL for all samples throughout theexperiment). After this time, all metal concentrations reduced sig-nificantly and stabilised. By the 2009/10 season, all treatments hadconcentrations of Cd, Cu, Fe, Mn, Sb, Se, and Zn near or below back-ground levels. The most effective fixation occurred for Mn and Zn,with final leachates below background and ANZECC/ARMCANZ(2000) guideline values. Concentrations of Co and Pb did not reachbackground levels but were decreased significantly below those inthe untreated landfill. The only increase in concentration betweenseasons was recorded for Fe, however, levels were still belowuntreated landfill and background values.

Although As and Cr concentrations in the treated materialswere not reduced below untreated landfill or background controlvalues, there was a greater than 10-fold decrease from the 2008/09 maxima to the stable 2009/10 concentrations. Phosphate fixa-tion was least successful for Ni. At the end of the experiment in2009/10, leachate Ni concentrations were up to an order of magni-tude greater in the treated material compared to the untreatedlandfill.

The major elements P, Ca and Mg were present in the treatmentreagents, and hence at high concentration in the leachates. Calciumin the treated materials initially increased (but interestingly, didnot exceed concentrations leached from the untreated landfill),and then decreased to a uniform value between the untreatedlandfill and background levels by 2009/10. Similar trends occurredfor Mg and P, with both initially increasing in concentration in thetreated materials before decreasing and stabilising in 2009/10 (butat higher levels than background and untreated landfill controls).

In summary, all treatments increased metal concentrations,except for Fe and Sn, at the beginning of reactions in 2008/09, pro-ducing a significant contaminant pulse. However, the higher ratio(more Emag-buffered) treatments (D5 and D6) showed smaller ini-tial increases (�50%) compared to the lower ratio (less Emag-buffered) treatments (D3 and D4) for As, Cd, Co, Cr, Cu, Mn, Ni, P,Sb, and Zn. Subsequently, all treatments reduced the leachate con-centrations of many of the contaminant metals, demonstrating theefficacy of the fixation technique after �800 h. The most effectivefixation occurred for Mn and Zn and the most effective treatmentacross all drums and for most metals was the 3:2 ratio of TSPand Emag (D5). The least effective treatment was the 2:1:1 ratioof TSP, Emag and phosphate rock (D4).

4. Discussion

Testing phosphate fixation in cold environments provides vitalinformation on the effectiveness of this remediation technique onlandfill materials in these locations. This Antarctic study ofmetal-phosphate fixation was carried out on landfill material

contaminated mainly with Cu, Pb and Zn, and allowed to undergo14 months of natural freeze–thaw cycles.

4.1. Environmental and chemical factors controlling treatment efficacy

Although the goal of phosphate amendment is to maximisemetal fixation, minimising phosphate leaching to the environmentis also important to reduce the risk of eutrophication of receivingwaters. Treatments that only contained TSP buffered with Emag(D3, D5) were more successful at reducing metal concentrationsin leachates compared with treatments that also contained phos-phate rock (D4, D6). This was consistent with previous studieswhere applications of >10% w/w phosphate rock have been recom-mended due to its low solubility (Basta et al., 2001). Our studydemonstrated that the best treatment for the Thala Valley landfillmaterial was TSP and Emag at 3:2 ratio (D5), which resulted inthe greatest degree of fixation and, less than 100 h after the startof the experiment, consistently the lowest phosphate concentra-tion in leachates.

During past summers, a seasonal pulse of contaminants in melt-water from Thala Valley landfill flowed into Brown Bay (Snapeet al., 2001b). To simulate this summer melt, water was added atthe start of the experiment in 2008/09 (Fig. 5). Addition of thewater increased leachate concentrations of most elements overthe first 24–48 h, with metals mobilised by dissolution or desorp-tion, followed by advective transport in the solution phase or viaparticle sorption and entrainment. This initial increase was likelya consequence of the acidifying effect of phosphate treatmentsand the slow kinetics of fixation (including slow dissolution ofthe phosphate source) (Fig. 7).

The initial increase in leachate elemental concentrations wasmatched by a decrease in pH, with pH being less acidic for thehigher treatment ratios. Lowest pH was recorded for treatmentD4 (TSP:Emag:PR – 2:1:1; pH 5.09–5.70), and D4 also showedthe greatest increase in leachate metal concentrations, both ini-tially over the first 24–48 h and overall. In contrast, the higher ratiotreatments (D5 and D6) had the smallest initial increases in metalconcentrations. This is likely an effect of the Emag – the highertreatment ratios contained more Emag relative to phosphate, buf-fering the acidity of the phosphate and reducing metal dissolution.

The presence of oxidising conditions, the high proportion offines, and the dissolution of soluble salts in the landfill materialare also factors likely to have contributed to the production ofthe contaminant pulse observed for the treatments in the veryearly stage of the experiment.

While the reduction of iron or manganese oxyhydroxide phasesin soil and sediment may lead to the release of adsorbed metalsinto solution (e.g. Chuan et al., 1996), in general, reducing condi-tions decrease dissolved metal concentrations (María-Cervanteset al., 2011), whereas oxidation favours solubilisation of metalsincluding Cd, Cu and Zn (Conesa et al., 2011; Du Laing et al.,2009; Hermann and Neumann-Mahlkau, 1985; Otero and Macías,2002). The initial increase in Eh and decrease in pH over the first24 h coincided with the observed increase in metal concentrations.As the experiment progressed, however, Eh measurements indicatethat conditions became more reducing (Fig. 6), and by the end ofthe 2008/09 season are likely to have favoured the onset of bacte-rial sulphate reduction (Connell and Patrick, 1968). Organic matter,present especially within the landfill material, would have pro-vided a suitable substrate for this process, and the detection of asulphide odour suggests that it was very likely occurring.

At Thala Valley, fine particles (<75 lm) comprised �30–40% ofthe landfill and contributed the major proportion of metal contam-ination in waters discharged from the site (Northcott et al., 2003).Many of these particles were likely colloidal (0.001–1 lm, e.g.naturally occurring ferric-coated clays), providing an effective

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40 E.G. Hafsteinsdóttir et al. / Applied Geochemistry 51 (2014) 33–43

mechanism for the aqueous transport of contaminants from thesoil. During sampling of leachates, much of this material may havebeen incorporated in the <0.45 lm ‘dissolved’ fraction. Further-more, oxidising conditions can increase the concentration of col-loids and hence of colloid-bound metals, while lowering pHincreases the proportion of very small colloids (0.03–0.07 lm)(Calmano et al., 2005).

Electrical conductivities of leachates were high in both treatedand untreated landfill materials compared with the backgroundcontrol. Soils with leachates > 4 mS/cm and pH > 7 are character-ised as saline (Pierzynski et al., 2005), which applies to this landfillmaterial (Fig. 6). The high EC in the treated landfill material is partlyattributable to the added phosphate (Peryea, 1991). However, sincethe EC values of untreated and treated landfill are similar, it is morelikely due to the close proximity of the low-lying site to the ocean(hence affected by sea spray and direct inundation) (Fig. 1), as wellas the presence of contaminant electrolytes. The quarry, fromwhere the background material was collected, is elevated andlocated 500 m upwind of the sea behind a hill, so it is partly shel-tered from wind-blown salts and unlikely to have been subject torecent flooding from the sea.

Metal availability in soil increases with salinity owing to com-petition for sorption sites by the major cations Na, Mg, Ca and K(Du Laing et al., 2009; Li et al., 2010). Under these conditions, met-als desorb. Sediments suspended in seawater indicated metaldesorption in the order Ni > Cd > Zn > Cu (Du Laing et al., 2009).Therefore the high salinity of the landfill material is likely to havebeen a factor contributing to the initial increase in leachateconcentrations, with desorbed metals entrained by the added meltwater as it moved through the soil.

4.2. Phosphate fixation efficacy for individual metals

Phosphate fixation has shown potential for the remediation ofmetals, including Cd, Cu, Fe, Mn, Pb and Zn in temperate settings(Cao et al., 2004; Harris and Lottermoser, 2006; Liu and Zhao,2007; Xu et al., 1994). In contrast, phosphate fixation in cold andfreezing environments remains understudied. The formation andstability of Cu-, Pb- and Zn-phosphates have been investigated inlaboratory experiments involving model single- and multi-metalsystems (Hafsteinsdóttir et al., 2011, 2013; White et al., 2012),and their results support the findings presented here that phos-phate amendment can effectively remediate metal contaminatedsites.

In temperate environments, phosphate addition transforms Pbto insoluble pyromorphite (Pb5(PO4)3(Cl, Br, F, OH)) (Cao et al.,2008; Nriagu, 1972, 1973; Porter et al., 2004; Zhang and Ryan,1998). The formation and stability of lead phosphates, in particularpyromorphite, has been demonstrated in laboratory experiments atlow temperatures and during freeze–thaw cycling (Hafsteinsdóttiret al., 2011, 2013; White et al., 2012). Complementing this work,the reduction of leachable Pb in Antarctic landfill material by phos-phate amendment has been confirmed in the field. Natural freeze–thaw cycles over 1 year and the heterogeneous nature of the soil didnot hinder the efficiency of fixation, although reaction rates in theenvironment differed significantly from those measured in the lab-oratory. For example, White et al. (2012), reported rapid (�2 min)formation of pyromorphite in the laboratory at 2 �C and 22 �C, butin the field, 1 month was required to reduce the Pb concentrationof the leachate below that of untreated landfill.

In the terrestrial soil environment, moisture, particle size,chemical speciation (e.g. determined by chloride and organic mat-ter content) and climate (particularly temperature) control fixationrate (Pierzynski et al., 2005). For instance, the chemical form of thesource contaminant (PbCl2 or PbSO4) is a key factor in the rate ofpyromorphite formation (Hafsteinsdóttir et al., 2011; Whiteet al., 2012). A high concentration of chloride slows the formationof pyromorphite, which is likely at the seaside Thala Valley, wherehigh concentrations of lead chloride minerals have been identifiedin the landfill (White et al., 2012). Similarly, organic Pb complexescan slow formation of pyromorphite, preventing reactions going tocompletion over 380 d (Hashimoto et al., 2009). Thus the presenceof organic matter in the landfill may have slowed the process of Pbfixation, but was insufficient to prevent it.

The success of Zn fixation in the field investigation concurs withthe findings of laboratory experiments. During 240 freeze–thawcycles in model systems, simulating 20 years at 30 cm depth inthe soil profile at Casey, Zn phosphate minerals (sodalite; Na6Zn6

(PO4)6�8H2O, chiral zincophosphate; Na12(Zn12P12O48)�12H2O, andhopeite; Zn3(PO4)2�4H2O) formed in abundance and typicallyremained stable (Hafsteinsdóttir et al., 2013). The efficacy ofphosphate treatment for Zn has also been demonstrated intemperate environments, where 2.5% or 5% w/w hydroxyapatite(Ca5(PO4)3OH) and 2% w/w TSP reduced Zn concentration in planttissues by 51–88% over 6–22.5 weeks (Brown et al., 2005; Knoxet al., 2003); 0.5% w/w TSP decreased earthworm (Eisenia fetida)bioavailable Zn by 38–60% over 8 d (Maenpaa et al., 2002); andover 99% of Zn2+ was removed from solution via 30 min sorptionby hydroxyapatite in vitro (Xu et al., 1994).

The stabilisation of Cu and formation of Cu phosphate mineralsrequires a lengthy reaction time (up to 5 months), both at temper-ate and cold temperatures (Ayati and Lundager Madsen, 2000;Hafsteinsdóttir et al., 2013; Qian et al., 2009). In laboratory exper-iments, formation of Cu phosphate minerals during multiplefreeze–thaw cycles was difficult to confirm, particularly in amulti-metal system where greater amounts of the phosphatesource might be required to increase their abundance

E.G. Hafsteinsdóttir et al. / Applied Geochemistry 51 (2014) 33–43 41

(Hafsteinsdóttir et al., 2013). White et al. (2012) reported fasterand more complete Cu phosphate reactions at 2 �C compared to22 �C in model systems, reaching equilibrium in 10 d. This is inagreement with our findings, with Cu concentrations in leachatefrom the treated landfill material reduced to a level similar to thatfound in the untreated and background controls in �1 month.

Significant reductions in Co concentrations have been reportedusing phosphate fixation (Kaplan and Knox, 2004; Seaman et al.,2001; Suzuki et al., 1982). Knox et al. (2003) reported �80% reduc-tion in aqueous Co(II) in one week by hydroxyapatite in a temper-ate setting. In the cold environment studied here, treatment D5reduced Co leachable concentrations most successfully. All treat-ments, except D4, reduced Co concentrations below untreatedlandfill levels by 2009/10. Eh and pH were likely the main factorscontrolling Co concentration, with neutral pH and strongly reduc-ing conditions favouring formation of Co sulphides and decreasingCo solubility (McBride, 1994). These conditions were observed inall systems during 2009/10 sample collection.

Phosphate fixation was particularly effective for Mn. Manga-nese solubility increases under reducing conditions (Du Lainget al., 2009), and so an increase in Mn concentration with the onsetof reducing conditions was expected for all systems but was onlyapparent in the untreated material (Fig. 2S (Supporting Informa-tion)). All treatments except D4 had leachable Mn at or belowthe level in the untreated landfill control after �18 days, and bythe 2009/10 season, all treated materials exhibited stable Mn con-centrations below background (D2) Based on the Eh-pH conditionsof the treated material, the dominant P species were H2PO4

� andHPO4

2�, likely forming Mn3(PO4)2 or MnHPO4 according to equilib-rium predictions (Lindsay, 1979). Thermodynamic modelling hasshown that low apatite concentrations are sufficient to form thestable MnHPO4 (Porter et al., 2004). Studies in temperate environ-ments also show decreases in solubility of Mn by hydroxyapatite(Seaman et al., 2001; Suzuki et al., 1982).

Phosphate fixation of Cd in temperate environments is con-trolled by Eh-pH conditions rather than the type of phosphatesource (Matusik et al., 2008; Porter et al., 2004). However, the typeof phosphate source can affect pH and therefore the efficiency offixation (Thawornchaisit and Polprasert, 2009). Treatment D5,which best reduced Cd concentrations, produced a pH range simi-lar to that reported as most effective for Cd fixation (pH 6.75–9.00;Matusik et al., 2008). Cadmium sulphide (CdS) and cadmium phos-phate (Cd3(PO4)2) are the two forms of Cd generally immobile frompH 3–12 (Porter et al., 2004). A likely scenario for the Thala Valleylandfill material involves the successful fixation of Cd, hereassumed to be largely in the form of the weathering product CdS,through the favourable combination of Eh, pH and EC conditions.Under oxidising conditions, as at the start of the experiment(Fig. 6), Cd concentrations increased as CdS oxidised to the moresoluble cadmium sulphate (CdSO4), leaving phosphate as the pre-ferred salt for stabilisation (Hermann and Neumann-Mahlkau,1985; Porter et al., 2004). Oxidation occurs to a greater extent insaline soils such as the Thala Valley landfill (Fig. 6), resulting in fas-ter release of Cd from CdS (Du Laing et al., 2009) and enhancingcadmium phosphate formation. As redox conditions became morereducing and pH increased, leachable Cd concentrations decreased.

Previous investigation identified Fe-sulphate hydroxide(Fe(SO4)OH), and Fe-oxide (Fe3O4), with minor akaganeite(FeO(OH)) as the main Fe phases derived from the weatheringof macro-contaminants in the original Thala Valley landfillwaste stockpile (White et al., 2012). Visual inspection of thelandfill material and the initial positive Eh values confirmedthat oxidation of Fe containing materials had occurred. Achange to reducing conditions during the experiment, and theacidity created by the addition of phosphate treatments, likelycontributed to the significant increase in leachable Fe

concentration via solubilisation of these main Fe phases (DuLaing et al., 2009). As pH approached neutral, Fe remained insolution owing to the reducing conditions (Du Laing et al.,2009; Hermann and Neumann-Mahlkau, 1985). However, theleachable Fe concentrations in the treated materials are muchlower than in the untreated landfill control, which continuedto increase through to 2009/10. Phosphate binds strongly withFe(III) in soil, typically in combination with Al, forming a rangeof insoluble Al–Fe–PO4 minerals (Haseman et al., 1950;McBride, 1994). This suggests that phosphate additionsuccessfully countered the potential for Fe to solubilise in thelandfill – particularly important given the dominance of Fecontamination in the landfill, and the significant potential forFe-colloids to transport adsorbed contaminant metals in melt-water run-off.

Fixation was least successful for As, Cr and Ni. In temperate envi-ronments, the solubility of As and Cr increases with phosphateaddition (Peryea, 1991; Seaman et al., 2001). Arsenic solubility isenhanced via competitive phosphate–arsenate (PO4

3�–AsO43�) inter-

action (Peryea, 1991) and under reducing conditions (Hermann andNeumann-Mahlkau, 1985). Both of these factors are likely to havecontributed to the increased leachable As in the phosphate-treatedmaterial in 2008/09. The decrease in As in leachate from the treatedlandfill material into 2009/10 might be a consequence of higher Feconcentrations, as Fe-containing material (especially colloids) canenhance As adsorption (Kumpiene et al., 2008). For Cr, leachableconcentrations may have increased due to competitive effectsbetween the oxyanion species of Cr and PO4 (Seaman et al., 2001).Chromium solubility is also enhanced by salinity (Du Laing et al.,2009) (Fig. 6). Similarly, unsuccessful fixation of Ni might be attrib-utable to the high salinity of the landfill material (Fig. 6). In salinesediments more Ni desorption occurs compared to Cd, Zn, and Cu(Du Laing et al., 2009). Also, the reducing conditions in the experi-ment may have increased Ni mobilisation, as has been noted for Feand Mn oxides (Du Laing et al., 2007).

4.3. Phosphate fixation complemented by other remediationtechniques

The results of this investigation indicate that the efficiency ofphosphate fixation varies with the contaminant under treatment.Generally, increasing phosphate concentration to excess in fieldsituations to improve fixation efficiency is not recommended,owing to the risk of significant input of phosphate into the receiv-ing environment. This is illustrated clearly in the field experiment,with P concentrations in the second season (2009/10) around anorder of magnitude higher in the treated materials compared withthe untreated landfill and background controls. To help prevent therelease of excess phosphate into the receiving environment, anaccurate mass-balance estimate of reagents versus fixation prod-ucts is required.

Phosphate fixation for remediation of As, Cr, and Ni contami-nated materials is not recommended unless the technique isrequired for stabilisation of co-occurring metals. When used onmaterial containing a mixture of contaminant metals (e.g. landfills),other techniques should be employed concurrently to minimise theenvironmental risk caused through mobilisation of As, Cr, and Ni.For example, adsorbents such as zero-valent iron, which adsorbAs, Cr and Ni, could be used in conjunction with phosphate fixation(Kumpiene et al., 2007, 2008). Alternatively, a ‘treatment train’approach can be applied, where phosphate fixation is coupled withdownstream water treatment systems (Northcott et al., 2007) orpermeable reactive barriers (PRBs) (Gore, 2009). PRBs provide aparticularly attractive cost-effective option as they perform wellin remote environments owing to ease of installation, and requireminimal infrastructure, monitoring and maintenance, minimal to

42 E.G. Hafsteinsdóttir et al. / Applied Geochemistry 51 (2014) 33–43

no energy input, and are suitable for variable temperature andwater flow conditions (Snape et al., 2001a). PRBs can hold varioustreatment materials, depending on the type of contaminant beingmanaged, and be sequenced to cope with a diversity of contaminanttypes (Gore, 2009; Mumford et al., 2013). They can also be used tocapture excess treatment reagents to minimise the risk of contam-inating receiving waters (e.g., phosphate and eutrophication).Therefore, if off-site dispersal of contaminants can be controlled(Tin et al., 2009), a low-cost remediation technique such as phos-phate fixation offers an effective in situ strategy for the cleanup ofmetal contaminants in soil.

5. Conclusions

The most successful phosphate treatment for the Thala Valleylandfill material was TSP and Emag at 3:2 ratio (total 5% w/w), bothfor reducing metal concentrations and excess phosphate in theleachate. Although there was an undesirable initial flush of metalsfrom the contaminated soil in the first 24–48 h after treatmentaddition, leachable concentrations were reduced and stabilisedby the second season. Relative to levels in the untreated landfillmaterial, leachable concentrations of Cd, Cu, Co, Fe, Mn, Pb andZn were decreased by phosphate addition, whereas the techniqueincreased concentrations of As, Cr and Ni. During a full-scale fieldimplementation, a complementary technique such as a permeablereactive barrier, would contain or treat contaminated runoff untilleachates reduced to acceptable concentrations.

Eh, pH and EC play important roles in determining fixation effi-cacy. High EC values were mainly attributed to the proximity of thelandfill to the ocean, not the added phosphate. High salinity canincrease the mobility of metals in soil. An initial decrease in pHand increase in Eh for leachates was associated with a concomitantincrease in leachable contaminant concentrations. The fall in pHwas, at least in part, attributed to the dissolution of the phosphatetreatments, whereas the rise in Eh was mainly ascribed to oxygen-bearing meltwater moving through the soil. As the reactions con-tinued, systems became less acidic and more reducing, decreasingleachable metal concentrations.

Although reaction times are relatively long in this cold andfreezing environment, phosphate remediation has been shown tobe successful for a number of important metal contaminants foundin contaminated sites in remote regions with freezing ground.

Acknowledgments

We thank the Australian Antarctic Division (AAS2937), and theAustralian Research Council, PANalytical and Veolia EnvironmentalServices for financial support under grant LP0776373. Specialthanks to Tim Spedding, Meenakshi Arora and Trevor Bailey forfield support and Greg Hince for help with ICP-OES analyses.

Appendix A. Supplementary material

Supplementary data associated with this article can be found, inthe online version, at http://dx.doi.org/10.1016/j.apgeochem.2014.08.011.

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