Science of the Total Environment 490 (2014) 659–670

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Recent history of sediment metal contamination in Lake Macquarie,Australia, and an assessment of ash handling procedure effectiveness inmitigating metal contamination from coal-fired power stations

Larissa Schneider a,⁎, William Maher a, Jaimie Potts b, Bernd Gruber a, Graeme Batley c, Anne Taylor a,Anthony Chariton c, Frank Krikowa a, Atun Zawadzki d, Henk Heijnis d

a Institute for Applied Ecology, University of Canberra, Canberra, ACT 2601, Australiab New South Wales Office of Environmental and Heritage, Lidcombe, NSW, 2141 Australiac CSIRO Land and Water, Lucas Heights, NSW 2234, Australiad Institute for Environmental Research, Australian Nuclear Science and Technology Organisation, Lucas Heights, NSW 2234, Australia

H I G H L I G H T S

• The main sources of metals to Southern Lake Macquarie are coal-fired power stations.• The metal of highest concern in this estuary is cadmium.• Arsenic was mobile in sediments.• Selenium and cadmium decreased in sediments following new ash handling procedures.

⁎ Corresponding author. Tel.: +61 429088813.E-mail address: Larissa.Schneider@canberra.edu.au (L.

http://dx.doi.org/10.1016/j.scitotenv.2014.04.0550048-9697/© 2014 Elsevier B.V. All rights reserved.

a b s t r a c t

a r t i c l e i n f o

Article history:Received 22 December 2013Received in revised form 9 April 2014Accepted 15 April 2014Available online xxxx

Editor: F.M. Tack

Keywords:CoalPower stationAsh dam137Cs210PbSediment

This study assessed historical changes inmetal concentrations in sediments of southern LakeMacquarie resultingfrom the activities of coal-fired power stations, using a multi-proxy approach which combines 210Pb, 137Cs andmetal concentrations in sediment cores. Metal concentrations in the lake were on average, Zn: 67 mg/kg, Cu:15 mg/kg, As: 8 mg/kg, Se: 2 mg/kg, Cd: 1.5 mg/kg, Pb: 8 mg/kg with a maximum of Zn: 280 mg/kg, Cu: 80mg/kg, As: 21 mg/kg, Se: 5 mg/kg, Cd: 4 mg/kg, Pb: 48 mg/kg. The ratios of measured concentrations in sedimentcores to their sediment guidelines were Cd 1.8, As 1.0, Cu 0.5, Pb 0.2 and Zn 0.2, with the highest concern being forcadmium. Of special interest was assessment of the effects of changes in ash handling procedures by the ValesPoint power station on the metal concentrations in the sediments. Comparing sediment layers before and afterash handling procedures were implemented, zinc concentrations have decreased 10%, arsenic 37%, selenium20%, cadmium 38% and lead 14%. An analysis of contaminant depth profiles showed that, after implementationof new ash handling procedures in 1995, selenium and cadmium, the main contaminants in Australian blackcoal had decreased significantly in this estuary.

© 2014 Elsevier B.V. All rights reserved.

1. Introduction

Coastlines are some of the most highly developed regions of theworld as these areas are very favourable for trade, communication andmarine resource exploitation (World Resources Institute, 2013). This de-velopment, as a consequence, has caused the disruption of the ecologicalintegrity of marine and estuarine lake ecosystems (Roy et al., 2001).

Estuaries are among themost important coastal features in Australia,both ecologically and with respect to human settlement and use (Ryanet al., 2003). In New South Wales, a state on the east coast of Australia

Schneider).

(Fig. 1), there are approximately 154 large and medium-sized estuariesand embayments (Butler and Jernikoff, 1999). LakeMacquarie, the larg-est of these estuaries, has been a strategic location for coal-fired powerstations (Batley, 1987). First, the existence of major underground coaldeposits has provided the power stations with a reliable fuel sourcefor generating electricity. Second, Lake Macquarie waters have beenused for generating steam to drive steam turbines and for cooling theexhaust steam from these coal-fired power stations. Third, thesepower stations are located in close proximity to Newcastle, a heavy in-dustrial area, where the electricity generated by the power stations isreadily used.

Despite the industrial development and prosperity these coastalareas have offered toNSW, the release of contaminants includingmetals

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andmetalloids into the environment through runoff has caused a rangeof environmental issues (Batley, 1987, 1992). The intensive industrialactivity on this estuary has resulted in fish having selenium concentra-tions above those recommended for human consumption (Roberts,1994; Kirby et al., 2001; Roach, 2005). Subsequently bans on commer-cial and recreational fishing in this estuary were imposed (Jackson,1995; Woodford, 1995).

Tomitigate the problem ofmetal release into the Lake, in 1995 ValesPoint Power Station implemented a different mechanism of coal ashhandling at its ash dam (Carroll, 1999). Coal ash used to be pumped asa slurry along a pipeline to the ash dam, where solids settled out andwater levels were regulated by gravity flow. During periods of highwater levels in the dam, the overflow water used to be drained intoLake Macquarie. Vales Point Power Station has adopted an operationin 1995 where the overflow water from Mannering Lake ash dam isrecycled back to the station under normal operating conditions. Onlyduring wet weather periods and pump stoppages does the effluentdischarge into Lake Macquarie and, therefore, metal contamination isexpected to have reduced in this estuary since 1995 (Carroll, 1999).

Fig. 1.Map of LakeMacquarie and coal-fired power station locations. Sediment core sites are idethe locationwhere thewater from the ash dams enters into the lake:White Heads Lagoon (Erarilocation of Lake Macquarie in New South Wales, Australia.

In this study, we investigated metal accumulation in sediments inSouthern Lake Macquarie over the past 60 years, which encompassesthe time coal-fired power stations have operated. This study soughtto: i) assess historical changes in sediment metal concentrations usinga multi-proxy approach which combines 210Pb dating and the,measurement of 137Cs and metal concentrations in sediment cores;ii) understand spatial metal concentration changes in sediments overtime per location; and iii) test the hypothesis that metal concentrationshave decreased after changes in coal ash handling procedures wereimplemented in 1995 by Vales Point Power Station.

2. Materials and methods

2.1. Historical setting

LakeMacquarie is situated in the lower Hunter Valley, approximate-ly 90 kmnorth of Sydney and close to the city of Newcastle (Fig. 1). Thisestuary extends approximately 22 km in a north–south direction, has amaximum width of about 10 km, a maximum depth of approximately

ntified by circles. Asterisks indicate cores selected for dating analysis. Black arrows indicateng Power Station);Mannering Bay (Vales Point Power Station). On the top left corner is the

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11m,with an average depth of 8m (Maunsell and Partners, 1974). Thisestuary is separated from the ocean by a narrow entrance channel andsand-bars at Swansea. The tidal range in Lake Macquarie is small withthe spring tidal range being estimated as 0.15 m at the western end ofthe tidal channel (Roy and Peat, 1976), decreasing with distance fromthe entrance to 0.06 m. Despite this poor tidal exchange, this estuaryhas a marine character because of minimal freshwater dilution fromthe two main fluvial inputs. Tidal currents are non-existent in most ofthe estuary, with the exception of the tidal channel at Swansea; windsare considered to produce larger changes in water levels inthe estuary than do tides (Roy et al., 2001). Shallow waters betweenSwansea and Wangi Wangi Point effectively bar deepwater movementwithin the estuary, resulting in a division of the estuary into northernand southern components about this latitudinal axis. As such, watermovements in the two portions of the estuary are essentially indepen-dent (Spencer, 1959).

In the northern part of the estuary, industrial development wasextensive, with a lead–zinc smelter, a fertiliser plant, a steel foundry,collieries and sewage treatment works operating in the past (Batley,1987). Metals released by these industries into the estuary are not be-lieved to be transported to the southern part because the two portionsof the estuary are essentially independent. A lead–zinc smelter operatedfor over 100 years near Cockle Creek at the northern end of LakeMacquarie until its closure in 2003. This industry was the main sourceof lead, zinc and cadmium released into this part of the estuary(Batley, 1987). The smelter is located a long distance (20 km) fromthe power station and there is no evidence that it has been a significantsource of atmospheric releases of metals. All the smelter contributionshave been aqueous inputs and as they are adsorbed to particulatesand deposited, their transport does not extend to the south of the estu-ary. Apart from a sewage treatment plant in the north of the estuary,there are no metal discharges apart from the power stations that affectthe southern end of the estuary (Batley, 1987).

In the southern part of the estuary, electricity is currently generatedfrom burning coal at Eraring and Vales Point Power Stations, and waspreviously generated at Wangi Power Station. The power station atWangi Wangi was commissioned in 1958, with a capacity of 330 MWbut was decommissioned in 1986 because it was replaced with a moremodern station at Eraring. Vales Point started operations in 1963 andEraring Power Station in 1981. Both power stations are still operatingand have a combined output of 5390 MW, which is a large part ofNSW's generating capacity (Peters et al., 1999a). A change in metal re-leases to Lake Macquarie is purported to have occurred in 1995, whenVales Point instigated new ash handling practices (Peters et al.,1999b). Previously, the overflow water from the ash dam used to bedrained into Lake Macquarie. The practice of recycling of the ash damoverflow discharge to be used in the cooling water system may be ex-pected to reduce the amount of suspended selenium reaching the estu-ary (Carroll, 1999).

In addition, copper and zinc have been shown to be released fromcooling water piping in power stations (Batley, 1987). Although studieshave shown that copper is one of the most frequently measured metalsin highway runoff and increased development could create additionalmetal sources (NCHRP, 2004), there are no major roadways in proxim-ity to the estuary that would significantly affect the sites studied.

2.2. Sediment collection

Sediment cores were collected from nine sites in Lake Macquarie,chosen based on their proximity to the power plants. Two replicatecores (0.1 m distance apart) were collected per site in order to obtainaverage profiles and check that the sediment profile was not disturbedby collection procedures or by local human activities. Three cores (atsites M7, M8 and M9, Fig. 1) in Lake Macquarie did not have replicatesand were only used for dating analysis.

Sediment cores were collected using acid-washed polyvinylchloridetubes. After retrieval, the coresweremaintained in a vertical orientationand a steel plunger used to push the sediment up to the top part of thetube. The 15 cmdeep coreswere sliced every 3 cmusing a stainless steelspatula and ruler. Sediments were stored in zip-lock plastic bags at 4 °Cuntil analysis.

2.3. Metal analysis

Prior to metal analysis, all sediment core sections were freeze-driedfor 72 h. After drying, samples were placed into 200 mL tubes and ho-mogenization of samples was performed by intensive manual mixingof sediment. Approximately 0.2 g of freeze-dried material was weighedinto a 60 mL polytetrafluoroacetate (PFA) closed digestion vessel (MarsExpress) and 2 mL of concentrated nitric acid (Aristar, BDH, Australia)and 1 mL of 30% concentrated hydrochloric acid (Merck Suprapur,Germany) added (Telford et al., 2008). Each PFA vessel was thencapped, placed into a 800 W microwave oven (CEM model MDS-81,Indian Trail, NC, USA), and samples heated at 120 °C for at least15min. The diluted digests were cooled to room temperature and dilut-ed to 50mLwith deionisedwater (Sartorius). The tubes were then cen-trifuged at 5000 rpm for 10 min. One millilitre of the digest wastransferred into a 10-mL centrifuge tube and then diluted to 10 mLwith ICP-MS internal standard (Li6, Y19, Se45, Rh103, In116, Tb159,Ho165). Digests were stored (0–5 °C) until analysis.

Samples were analysed for metals using inductively coupled plasmamass spectrometry (ICP-MS) (Perkin-Elmer DRC-e)with anAS-90 auto-sampler (Maher et al., 2001). Selenium concentrations were cross-checked on a Perkin-Elmer 5100 Zeeman graphite furnace atomicabsorption spectrometer, with an AS-60 auto-sampler, using a mixtureof palladium and magnesium as the matrix modifier (Deaker andMaher, 1995).

The certified reference materials BCSS-1 and MESS-1 from the Na-tional Research Council Canada and 1646a from the National Instituteof Standards and Technology were used as controls to check the qualityand traceability of metals. Measured concentrations were in agreementwith certified values (Supplementary Table 1).

2.4. Dating analysis

The sediments were dated using 137Cs and 210Pb. The total 210Pbactivity in sediments has two components: supported 210Pb activity(210Pbs) which derives from the decay of in situ radium-226 (226Ra),and unsupported 210Pb activity (210Pbu)whichderives fromatmospher-ic fallout. 210Pbs activity is determined indirectly via its grandparentradioisotope 226Ra. 210Pbu activity was calculated by subtracting 210Pbsactivity from the total activity of 210Pb (210Pbt) or its progenypolonium-210 (210Po).

The in-situ production of supported 210Pb was measured indirectlyfrom the activity of 226Ra using gamma spectrometry. Unsupported210Pbu cannot be measured directly and so is inferred from the activityof total 210Pbt minus the activity of supported 210Pbs. The activity oftotal 210Pbt was determined by measuring 210Pb directly using gammaspectrometry.

Between 20 and 40 g of dried and ground whole sediment wasanalysed for 137Cs, 210Pb and 226Ra activities by Compton suppressiongamma spectrometry. The detector system energy calibration was car-ried out using a National Institute of Standards and Technology (NIST)traceable multi-nuclide standard source (QSA Global no. 121881-RF929) and the detector system efficiency calibration was determinedusing IAEA reference materials including RGU-1, RGTh-1, RGK-1 andSoil-6.

Both the Constant Initial Concentration (CIC) and Constant Rate ofSupply (CRS) models (Appleby and Oldfield, 1992) of 210Pb datingwere used to construct chronologies for each of the sites, using the137Cs date as a reference point. Inflections in the 210Pbu, activity profile

Table 1Whole sediment metal concentration (mg/kg) and normalized metal concentration by depth of cores from locations 1 to 9.

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Cell highlighted in grey indicates that metal concentrations were normalized to grain size. Cells highlighted in orange indicate metal concentrations above the Australia New ZealandGuidelines for Fresh andMarineWater Quality (ANZECC/ARMCANZ, 2000). Cells that were not highlighted indicate that sample concentrationswere below guidelines. Note that seleniumis not included in the guidelines (highlighted in blue).

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usually reflect changes in sediment accumulation rate or bioturbation(Krantzberg, 1985). For those sites where the weight accumulationrate is fairly constant, i.e. where 210Pbu, activitymonotonically decreaseswith depth, the CIC and CRS models are usually in agreement. Wherethere is a non-monotonic trend in the 210Pbu, depth profile, the use ofthe CIC model is usually precluded and the CRS model is used to calcu-late the chronology.

2.5. Grain size analysis

All sections of the cores were analysed for sediment grain size byfirst sieving to the particle size of b2000 μm (i.e. gravel-free) and, sedi-ment fractions were determined using laser detection on a MalvernMastersizer 2000 with a 300RF lens, 0.05–900. Samples were dispersedin water and 30 second ultrasonication used to break up agglomeratedparticles. A total of threemeasurementsweremade for each sample andthe average used as the final value. The distribution abundances withinthe 0.2 to 2000 μm grain size range were calculated by the MalvernMastersizer software.

All sample sediment sizes were classified to four classes: clay (parti-cles with diameter b2 μm), silt (2–20 μm), fine sand (20–200 μm), andcoarse sand (200–2000 μm) as recommended byMcKenzie et al. (2000)and Carlile et al. (2001).

As surface areas of sediments are grain-size dependent and controlthe adsorption of metals in sediments, metal concentrations werenormalized to grain size in order to interpret metal concentrations insediments between sites (Blomqvist et al., 1992; Hanson et al., 1993).

In this study, normalization of metal concentrations in sedimentswas achieved using iron following the methodology of Suh and Birch(2005). Normalized and non-normalized metal concentrations are re-ported in Table 1. The correlation between iron concentration and finefraction (clay) was the basis for the normalization. Clay content(b20 μm) significantly predicted iron concentration in total sediment(grain size b 2000 μm) (slope = 225.2, t (33) = 4.14, p b 0.001). Claycontent percentage explained 34% of variance in iron concentrationsin total sediments R2 = 0.34, F(1,33) = 17.22; p b 0.001 (Supplemen-tary Fig. 1). Iron concentrations were corrected to the clay content byusing the regression Eq. (1):

Normalized Fe ¼ 8820þ 225� clay contentð Þ: ð1Þ

Metal concentrations in sediments were normalized by multiplyinga given metal concentration value by 10,000 and dividing by the iron-normalized value. Normalized data were only used for comparisons ofmetal concentrations between sites. As grain size did not significantlychange vertically within cores, the non-normalized data were usedwhen dating data was used to interpret sediment metal concentrations.

2.6. Statistical analysis

To compare the pattern of metal concentrations by sites and bysediment depths, a dissimilarity matrix was constructed and variableswere tested using an Analysis of Similarities Test (ANOSIM) (Clarke,1993). ANOSIM explicitly compares the level of similarity found in allpairs within each level of a factor against the similarity found inbetween-level pairs. Significance testing is obtained through randomi-zations (Legendre and Legendre, 2012).

Specifically, this procedure was used to test the null hypothesis thatthere is no difference in metal concentrations in sediments betweensites and within depth. If the p-value was less than 0.05, then the nullhypothesis was rejected and a t-test pairwise with Bonferroni correc-tion was conducted at the level 0.05 based on dissimilarity ranks.

An ordination, Multidimensional Scaling (MDS), was used to inputthe resultingmetal concentration similaritymatrix. Each sitewas placedin the MDS algorithm in N-dimensional space such that the between-object distances were preserved.

All computations were performed using the software R 2.14.1(R Development Core Team, 2013), using the package Vegan (version2.0–7; URL: http://cran.r-project.org/web/packages/vegan/index.html).The graphs of profile cores were plotted using the Analogue package(version 0.10–0; http://cran.r-project.org/web/packages/analogue/index.html) and package Lattice (version 0.20–15 http://cran.r-project.org/web/packages/lattice/index.html).

3. Results

3.1. Metal concentration differences between sites

Metal concentrations in cores were different between sites (Fig. 2,Table 1). The metal concentrations (except arsenic) were relatively

Fig. 2. Profiles of normalizedmetal concentrationswithin sediment cores of LakeMacquarie by collection sites. Black dots represent the two replicate cores collectedper site and black linesrepresent the average metal concentration between replicates.

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higher at sites M1, M2, M3, and M5 than M4 and M6, with M6 consis-tently having lower metal concentrations.

Dissimilarity rankings of normalized metal concentrations insediment cores confirmed that significant differences in metalconcentrations occurred between sites (R = 0.33, p = 0.001; Fig. 3).Pairwise comparisons using the Bonferroni t-test revealed that sitesM1, M2 and M5 had the highest dissimilarity rank and had similarsediment metal concentrations. Sites M3 and M4 had an intermediatedissimilarity rank and shared similarities in metal concentrations. SiteM6 had the lowest dissimilarity rank, sharing similarities in metalconcentrations only with site M4 (Table 2, Fig. 3).

Sites M1, M2 and M3 were particularly high in all five metalconcentrations while M2 was particularly high in copper concentrations.M6, although located just next to the ash dam output of Eraring PowerStation, had lower metal concentrations than other sites. Representationof the data in aMDS plot (Fig. 4) also shows the differences between sitesconfirmed by the ANOSIM. M5 and M1 were spread, with some sampleshigher in arsenic, lead, zinc and cadmium concentrations.

3.2. Within core metal concentration variation with depth (time)

Metal concentrations over depth (time) were very similar withinsites, with the exception of arsenic (Fig. 2). Arsenic concentrations hada unique behaviour with depth, with different trends in concentrationchanges from other metals.

Based on metal background concentrations previously reportedfor Lake Macquarie of zinc b 70 mg/kg, copper b 20 mg/g, arsenicb 15 mg/g, selenium b 1 mg/g, cadmium b 0.8 and lead b 15 mg/kg(Roy and Crawford, 1984; Kilby and Batley, 1993; Peters et al., 1997,1999a), in this study, copper, zinc and selenium concentrations reachedbackground values at the bottom layers of the sediment cores at sitesM1 and M5 (Fig. 2). Arsenic reached background concentrations atsite M6, cadmium at site M5 and lead at sites M2, M5, and M6.

Unsupported 210Pb activities for all three cores used for datinganalyses (cores M7 to M9; Fig. 1) exhibited a decreasing profile withincreasing depth (Figs. 5, 6 and 7). The calculated CIC and CRS modelsediment ages were in agreement for all three cores (Table 3).

Fig. 3. Boxplots of ranked metal concentration dissimilarities of sediments from Lake Macquarie, NSW. Pairwise comparisons were grouped as between sites or within site (individualgroups shown in the plot). Boxes represent the median and interquartile range (IQR); whiskers extend to the most extreme data points up to 1.5 times the IQR; open dots representoutliers outside this range. Letters shared in common between or among the groups indicate no significant difference.

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The 137Cs profile could only be used to validate the 210Pb datingchronology in core M7, with 137Cs activities higher in the top 9 cm ofthe core, followed by a 137Cs activity decrease to 1.7 Bq/kg at 3–6 cm.According to the 210Pb chronology, the sediment layer at 3–6 cm is~43 years old, whereas the sediment at 6–9 cm is ~72 years old. CoresM8 and M9 had significant levels of 137Cs activity detected in all thesamples and could not therefore be used to validate the 210Pb datingchronology.

Based on 210Pb results, cores M7 and M8 had a constant sedimenta-tion rate of 0.1 cm/y while M9 had a sedimentation rate of 0.2 cm/y.Metal concentrations reached amaximumatmid depth (3 to 9 cm), cor-responding to the peak of inputs from power station activities (1980),and all metals studied (zinc to a lesser extent) decrease from thislayer to the surface showing that metals decreased since ash dam over-flowmodifications (Figs. 5 and 6). Copper, zinc and lead concentrationsdid not decrease significantly in the top layer 0–3 cm (Figs. 5 and 6).

For core M9, only selenium and cadmium concentrations slightlyfollowed the power station activity pattern (Fig. 7).

4. Discussion

4.1. Metal concentrations in sediments: difference between sites

Metal concentration variation between and within cores (Fig. 2)showed the importance of considering the geographic shape of the estu-ary and its water flow dynamics when studying metal distribution in

Table 2Pairwise comparisons of metal dissimilarity ranks between cores from Lake Macquarie.Asterisks are p values.

Sites Between M1 M2 M3 M4 M5

M1 1 – – – – –

M2 1 1 – – – –

M3 ⁎⁎⁎ ⁎⁎ 0.2 – – –

M4 ⁎⁎⁎ ⁎⁎⁎ ⁎⁎ 1 – –

M5 1 1 0.9 ⁎⁎⁎ ⁎⁎⁎ –

M6 ⁎⁎⁎ ⁎⁎⁎ ⁎⁎⁎ ⁎⁎ 0.3 ⁎⁎⁎

⁎⁎ p b 0.01.⁎⁎⁎ p b 0.001.

sediments. Cores M4 and M6 had the least changes in metal concentra-tions within the core. This is a result of their location in an open area ofthe estuary and water flow moving particles rather than letting themsettle in the area (Fig. 1). Eraring and Vales Point Power Stations pro-duce a similar amount of energy, but Vales Point ash dam is locatednext to Mannering Bay, a closed area which is historically known forits sediment to act as a depositional zone for contaminants (Batley,pers. comm.). This is confirmed by cores M2, M3 and M5 (Fig. 2) thatwere located in bays (Fig. 1) and reflect the greatest metal changes ofall the sites.

The appropriateness of collecting sediments in bays rather thanopen areas is also shown by the ANOSIM results. Sites M1, M2 and M5are characterised by strong heterogeneity between samples within asitewhileM3,M4 andM6have little variance (Fig. 4). Siteswith highestheterogeneity between samples correspond to the sites located in bayswhile the sites with the lowest variance, with the exception of M3, arethe sites located in open areas of the estuary. Cores located in openareas have more water flow and wind influences are more susceptibleto sedimentmixing resulting in a consistent pattern ofmetal concentra-tions within cores. Therefore, distance, geomorphology and hydrologydirectly determined the distribution of metals in sediments and, as forthis study, they should be considered in the experimental designwhen studying metal history in sediments.

Similar to other studies that have measured metal concentrations insediment ash damdischarges,we found elevated concentrations of cop-per, cadmium and selenium, however, unlike these studies we have notfound elevated concentrations of arsenic, mercury and cobalt. In LakeErie, the fourth largest lake of the five Great Lakes in North America, ar-senic and cobalt were found in greater concentrations in sediments lo-cated near an ash-dam compared with reference stations (Hatcheret al., 1992). In LakeWabamun, in Alberta, Canada, sediment concentra-tions of mercury, copper, lead, arsenic and selenium have increased by1.2 to 4 fold as the lake receives discharge from ash lagoons (Donahueet al., 2006) while in the Grand Lake, New Brunswick, Canada, Lalondeet al. (2011) reported a significant source of arsenic, thallium, andantimony to sediments located ≤100 m from the outfall but in thesediment of the East River, Nova Scotia, the ash dam was not a signifi-cant source of metals. The differences are a reflection of the differentcomposition of Australian coals which are not highly enriched in arsenicand mercury.

Fig. 4.MDS plot showing between-site similarity in metal concentrations in sediments from Lake Macquarie, NSW. Plot based on data for six original cores (circles) and their replicates(triangles), with points colour coded for sites. Arrows represent the relative contribution of the variables to the variability along the first two MDSs. (For interpretation of thereferences to colour in this figure legend, the reader is referred to the web version of this article.).

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In Australia, sediment quality guidelines(ANZECC/ARMCANZ, 2000),are used for assessing site contamination and determining the require-ments for further investigation. This document, advises further investi-gation if sediment concentrations are above the guideline trigger valuesof 1.5 mg/kg for cadmium, 65 mg/kg for copper, 50 mg/kg for lead and500 mg/kg for zinc. There is no guideline for selenium in Australiansediments as a result of limited data availability on selenium toxicity.

The sediment metals were ranked based on the ratio of the mea-sured concentrations to guideline values, yielding an order ofdecreasing environmental concern: Cd 1.8, As 1.0, Cu 0.5, Pb 0.2 andZn 0.2, indicating an order of ecological concern. Highest concern inLake Macquarie was for cadmium, which was above the guidelinetrigger value at all sites except M6 and M9.

Fig. 5. LakeMacquarie (CoreM7) depth profiles of a) grain size distributionswith depth; dark grcoarse sand (200–2000 μm), b) total concentration ofmetals and c) 210Pb activity depth profile.grey arrow corresponds to the periodwhen new ash handling procedureswere applied. Grey linARMCANZ, 2000). Lines on the extreme right side of the stratigraphy plot denote values that arfigure legend, the reader is referred to the web version of this article.).

4.2. Within sediment core metal concentration distribution

Within sites, the different metals had similar concentration profilesover time, except for arsenicwhich is amoremobile element. This resultalso shows that cores were not mixed during collection procedures andrepresent a reliable history of metal inputs.

The differences in variation in arsenic concentrations in sedimentcores compared to the other metals may be explained by mobilization,which is intimately tied to their oxidation state and redox reactions. Ar-senic is unique among the metalloids and oxyanion-forming elementsin its sensitivity to mobilization at the pH values of 6.5–8.5 under bothoxidising and reducing conditions (Smedley and Kinniburgh, 2002).While most metals occur in solution as cations which generally become

ey% clay (b2 μm),mediumgrey% silt (2–20 μm), grey% fine sand (20–200 μm), light grey%The black arrow on the right axis corresponds to the start of power station activity and lighte indicates background values and red line indicates the guideline trigger values (ANZECC/e outside the graph scale of samples. (For interpretation of the references to colour in this

Fig. 6. LakeMacquarie (CoreM8) depth profiles of a) grain size distributionswith depth; dark grey% clay (b2 μm),mediumgrey% silt (2–20 μm), grey% fine sand (20–200 μm), light grey%coarse sand (200–2000 μm), b) total concentration ofmetals and c) 210Pb activity depth profile. The black arrow on the right axis corresponds to the start of power station activity and lightgrey arrow corresponds to the periodwhen new ash handling procedureswere applied. Grey line indicates background values and red line indicates the guideline trigger values (ANZECC/ARMCANZ, 2000). Lines on the extreme right side of the stratigraphy plot denote values that are outside the graph scale of samples. (For interpretation of the references to colour in thisfigure legend, the reader is referred to the web version of this article.).

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increasingly insoluble as the pH increases, arsenic becomes less stronglysorbed as the pH increases. At near-neutral pH, the solubility ofmost metal cations is severely limited by precipitation as, orcoprecipitation with, an oxide, hydroxide, carbonate or phosphatemin-eral, or more likely by their strong adsorption to hydrous metal oxides,clay or organic matter. In contrast, most oxyanions including arsenatetend to becomes less strongly sorbed (Dzombak and Morel, 1990).Relative to the other oxyanion-forming elements, therefore, arsenic isdistinctive as the most mobile element over a wide range of redox con-ditions (Smedley and Kinniburgh, 2002).

Remobilization of selenium in contaminated sediments was report-ed by Bowie et al. (1996). This remobilization can cause porewater con-centrations to create a concentration driving-force and a mass flux ofselenium from sediments into the overlying water column. The samemechanisms would also create a downward diffusional mass selenium

Fig. 7. LakeMacquarie (CoreM9) depth profiles of a) grain size distributionswith depth; dark grcoarse sand (200–2000 μm), b) total concentration ofmetals and c) 210Pb activity depth profile.grey arrow corresponds to the periodwhen new ash handling procedureswere applied. Grey linARMCANZ, 2000). Lines on the extreme right side of the stratigraphy plot denote values that arfigure legend, the reader is referred to the web version of this article.).

flux. In deeper sediment, reduction and sorption mechanisms wouldthen immobilise the selenium (Bowie et al., 1996).

The changes in arsenic and selenium concentration down coresshow that these two elements need to be cross-checked with other ele-ments. In this study, selenium concentrations follow the overall patternof other metals despite fewer changes within core profiles compared toother metals. Changes in arsenic, however, are not similar to the pat-terns of other metal concentrations.

Although the length of cores collected in this study was only 15 cm,based on background values reported in previous studies, the corescollected in Lake Macquarie were shown to reach background metalvalues at that depth. In previous studies on metal concentrations inLake Macquarie using cores deeper than 50 cm (Roy and Crawford,1984; Batley and Lincoln-Smith, 1993; Kilby and Batley, 1993; Peterset al., 1999a; Kirby et al., 2001), background values were reported as

ey % clay (b2 μm),mediumgrey% silt (2–20 μm), grey %fine sand (20–200 μm) light grey %The black arrow on the right axis corresponds to the start of power station activity and lighte indicates background values and red line indicates the guideline trigger values (ANZECC/e outside the graph scale of samples. (For interpretation of the references to colour in this

Table 3137Cs, 210Pb dating, sedimentation rate (Mean± uncertainties) and sediment grain size composition (μm) of sediment cores 7, 8 and 9 by depth. Uncertainties are quoted at the 2 sigma level.

Core Depth Total Supported Unsupported 137Cs CIC age CRS age Sedimentation % Clay % Silt % Sand % Coarse sand

ID (cm) 210Pb (Bq/kg) 210Pb (Bq/kg) 210Pb (Bq/kg) (Bq/kg) (y) (y) (cm/y) b 2 μm 2-20 μm 20-200 μm 200-2000 μm

7 0–3 53 ± 5 15 ± 1 38 ± 5 2.1 ± 0.2 14 ± 15 13 ± 4 0.1 3 50 38 87 3–6 31 ± 3 16 ± 1 15 ± 3 1.7 ± 0.1 43 ± 19 36 ± 6 0.1 6 60 34 17 6–9 29 ± 4 15 ± 1 15 ± 4 0.8 ± 0.2 72 ± 26 68 ± 8 0.1 4 52 44 07 9–12 19 ± 4 18 ± 1 2 ± 4 b 0.6 100 ± 34 110 ± 10 0.1 5 56 38 17 12–15 16 ± 3 15 ± 1 2± 4 b 0.4 129 ± 42 174 ± 13 0.1 4 48 40 88 0–3 89 ± 6 18 ± 2 71 ± 7 2.2 ± 0.2 13 ± 3 8.2 ± 3 0.1 2 35 61 38 3–6 73 ± 6 16 ± 2 58 ± 7 2.6 ± 0.3 39 ± 15 28 ± 5 0.1 2 34 60 48 6–9 51 ± 4 18 ± 1 34 ± 4 2.8 ± 0.2 65 ± 19 58 ± 8 0.1 2 33 61 48 9–12 30 ± 3 16 ± 1 14 ± 3 1.2 ± 0.1 91 ± 23 102 ± 10 0.1 2 35 60 38 12–15 18 ± 4 16 ± 1 2 ± 4 0.5 ± 0.1 117 ± 28 179 ± 13 0.1 2 31 61 69 0–3 93 ± 8 17 ± 2 76 ± 8 2.3 ± 0.3 9 ± 3 8 ± 3 0.2 2 41 55 29 3–6 85 ± 5 18 ± 1 68 ± 6 3.2 ± 0.2 26 ± 9 24 ± 5 0.2 3 47 51 09 6–9 47 ± 6 17 ± 2 31 ± 6 3.0 ± 0.3 43 ± 6 43 ± 6 0.2 3 47 49 29 9–12 37 ± 4 15 ± 1 22 ± 4 2.2 ± 0.2 60 ± 11 61 ± 8 0.2 3 45 52 09 12–15 24 ± 4 15 ± 1 9 ± 5 1.2 ± 0.2 77 ± 13 79 ± 9 0.2 3 50 46 0

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zinc b 70 mg/kg, copper b 20 mg/kg, arsenic b 15 mg/kg, seleniumb 1 mg/kg, cadmium b 0.8 mg/kg and lead b 15 mg/kg. In this study,only zinc and lead concentrations in coreM9 did not reach these report-ed background values at 15 cm. Batley (1987) described lead concentra-tion distributions in LakeMacquarie as closely parallel to those for zinc,and perhaps their concentrations above background value in thebottom layer of core M9 are a result of atmospheric transport and pre-cipitation from the north (Roy and Crawford, 1984) or these metalsbeing discharged by sewage and urban runoff (Batley, 1987).

4.3. Dating and sedimentation rates

137Cs was successfully used as a proxy to validate 210Pb results onlyfor core M7. 137Cs as a chronomarker can be inaccurate in saline sedi-ments due to its mobility in the sediment profile (Beasley et al., 1982;Sholkovitz and Mann, 1984). On encountering seawater, a significantfraction of surface-bound 137Cs can be rapidly desorbed from sedimentparticles by ion-exchange with dissolved cations such as Na+ and K+

(Everett, 2009). As a consequence, desorption of 137Cs from sedimentgrains into pore water and diffusion into deeper sediment can distortthe sediment profile, potentially leading to a downward shift in thefirst onset horizon.

In this study, the problems associated with using 137Cs to validate210Pb age were overcome by correlating metal sediment profiles withhistorical development of coal-fired power stations. This solutionworked verywell, especially for cores collected close to the point sourceas metals were useful chronological indicators.

Sedimentation rates based on 210Pb results for cores collected in thisstudywere extremely low: 0.1 and 0.2 cm/y. These sedimentation ratesgenerated metal concentration profiles over time in agreement withhistorical inputs in the estuary. In addition, LakeMacquarie's low depo-sition rate is in agreement with previous studies describing the estuaryas an extremely low-energy depositional environment both at presentand in the geological past (Roy and Crawford, 1984; Peters et al.,1999a). Roy and Crawford (1984) collected cores in Lake Macquarie of60 and 100 cm depth and found a sedimentation rate ranging from 0.1to 0.8mm/y (mean 0.3 mm/y) using 14C dating, however, carbon datingis unreliable for recent sediments. Peters et al. (1999b) found similarrates of 0.1 cm/y at Mannering Bay and 0.1 and 0.6 cm/y at NordsWharf using 210Pb dating.

4.4. History of contamination

The distributions of copper, zinc, arsenic, selenium, cadmium andlead concentrations within the dated cores M7 and M8 (Figs. 5 and 6)were in agreement with historical events in the southern part of the es-tuary contributing to the inputs of metals. Concentrations of these

metals in cores M7 andM8 increased in the bottom of the cores and de-creased at the top of the cores. The pattern was the same at the bottomfor the core M9, but at the top there was an increase (Fig. 7). The in-crease in metal concentrations in the bottom of cores is in agreementwith the start of the power station activities from 1959, with peaks in1982when the third power station started its activity. Arsenic, however,had greater variability of concentration when compared to the othermetals, especially in coreM8 (Fig. 6). This study shows that before rely-ing on arsenic concentration as a proxy for a history ofmetal contamina-tion, it should be crosscheckedwith othermetals as it is themostmobilemetal in sediments over a wide range of redox conditions.

Of the previous studies of metals in the estuary, only Peters et al.(1999a) and Kirby et al. (2001) have assessed metal concentrationsafter ash handling procedures were implemented. The results ofPeters et al. (1999a) reflected only two years of new ash handling pro-cedures, while the study by Kirby et al. (2001), besides not including210Pb dating analysis, collected cores only at Nord's Wharf, the area inthe most southerly part of Lake Macquarie known to be least affectedby metals from power stations (Batley, 1987; Peters et al., 1999a).

The increases in metal concentrations in the top of core M9 mightreflect the distance from Vales Point Power Station (Figs. 1 and 7) andmetals discharged by this power station are not reaching sufficientconcentrations in that location to reflect the improvement in ashhandling procedures. This finding corroborates previous studies whichconcluded that the history of contamination and sedimentation cannotbe establishedwith confidence usingmetals at sites far from establishedpoint sources, primarily due to the diverse sources and dates of intro-duction of these metals (Payne et al., 1997). This limitation was clearlyhighlighted in core M9.

Comparison of the metal sediment core concentration profiles withprevious studies is only possible with results from Peters et al. (1999a),who collected cores close to Vales Point Power Station and conducted210Pb analyses. In their cores, selenium concentrations fell to backgroundvalues by 15 cm depth, with a single peak corresponding to the maxi-mum concentration at a depth of 5 cm. Based on the sedimentationrate of 0.1 cm/y reported by Peters et al. (1999a) at Mannering Bay andalso reported for cores M7, M8 and M9 in this study, the metal peak isat about 6 cm depth and in agreement with results found by Peterset al. (1999a) which corresponds to the 1990s.

4.5. Effects of changes in ash handling procedures

As a mitigation measure to decrease the release of metals to LakeMacquarie from the overflow water from Vales Point Power Stationash dam, at the end of 1995, Vales Point Power Station started recyclingwater from the ash dam overflow rather than discharging it straight toLake Macquarie (Carroll, 1999).

669L. Schneider et al. / Science of the Total Environment 490 (2014) 659–670

In this study, cores M7 and M8 were collected close to Vales PointPower station and were dated and analysed for metal changes withinsediments. The top 3 cm layer of the sediment (Figs. 5 and 6) representssediment concentrations after the 1990s and shows a uniform decreaseof arsenic, selenium, cadmium and lead concentrations but none forcopper. Comparing metal concentrations in cores M7 and M8 betweenlayers 3 to 6 cm (dated ~before 1995) and 0 to 3 cm (dated ~after1995), zinc concentrations have decreased 10%, arsenic 37%, selenium20%, cadmium 38% and lead 14%. This decrease shows that the newash handling procedures have been effective in decreasing runoff con-taining metals to Lake Macquarie.

This is thefirst study to evaluate long-term changes ofmetal concen-trations in sediments of southern Lake Macquarie after improved ashhandling procedures were adopted. The results from this study demon-strate that recycling water from ash-dams overflow is an efficient miti-gationmeasure to decrease the release ofmetals intowater bodies fromcoal-fired power stations.

5. Conclusions

We describe the history of contamination in sediments of southernLake Macquarie since the commencement of operation of three coal-fired power stations on its shores. Analysis of copper, zinc, arsenic, sele-nium, cadmium and lead concentrations in sediment cores has shownthat they are elevated, with concentrations above background valuesand cadmiumconcentrations above theANZECC/ARMCANZ (2000) sed-iment guideline trigger value. Metal concentration profiles within coreswere consistent, with the exception of the metalloid arsenic which be-haved differently within sediment from themetals because of its mobil-ity over a wide range of redox conditions. We confirmed that arsenicneeds to be cross-checked with other metals when studying the historyof metal contamination in marine sediments. The effect of proceduresfor handling ash dam waters have been assessed for the first time.Concentrations of selenium and cadmium, the main contaminants inAustralian black coal, decreased significantly in sediments from thetime the new ash handling procedures were implemented. These re-sults show that recycling water from ash dams is an efficient mitigationmeasure to control metal inputs to water bodies from coal-fired powerstations and should be applied to other coal-fired power stations in thecountry.

Supplementary data to this article can be found online at http://dx.doi.org/10.1016/j.scitotenv.2014.04.055.

Conflict of interest

None declared.

Acknowledgements

We thank NSW Office of Environment and Heritage for providingsubstantial field and logistical support, and the Australian Instituteof Nuclear Science and Engineering (AINSE, grant numberALNGRA11116) for providing funds and access to the national facilitiesat ANSTO to conduct dating analysis. L. Schneiderwasfinanced by an In-ternational Postgraduate Research Scholarship (IPRS) funded by theAustralian Government.

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