quantifying reduction in ecological risk in penrhyn estuary, sydney, australia, following...

Quantifying Reduction in Ecological Risk in Penrhyn Estuary,Sydney, Australia, Following Groundwater RemediationJames Hunt,y Gavin Birch,y and Michael Warne*zyEnvironmental Geology Group, School of Geosciences, University of Sydney, Sydney, New South Wales, AustraliazWater Quality and Aquatic Ecosystem Health, Department of Environment and Resource Management, Dutton Park, GPO Box 2454,Brisbane QLD 4001, Queensland, Australia

(Submitted 1 June 2010; Returned for Revision 13 July 2011; Accepted 20 April 2011)

ABSTRACTThe environmental risk associated with discharge of contaminated groundwater containing a complex mixture of at least 14

volatile chlorinated hydrocarbons (VCHs) to Penrhyn Estuary, Sydney, Australia has previously been assessed. That probabilistic

ecological risk assessment (ERA) was undertaken using surface water monitoring data from 2004 to 2005. Subsequently, in

2006, agroundwater remediation systemwas installed andcommissioned toprevent further dischargeofVCHs into theestuary.

The present study assessed the ecological risk posed to the estuary after 2006 to evaluate the success of the remediation system.

The ERA was undertaken using toxicity data derived from direct toxicity assessment (DTA) of preremediation contaminated

groundwater using indigenous species, exposure data from surface water monitoring between 2007 and 2008 and the joint

probability curve (JPC)methodology. The risk posedwasmeasured in 4 zones of the entire site: source area (2), tributary (2), the

inner estuary and outer estuary at high, low, and a combination of high and low tides. In the 2 source areas, risk decreased by

over 2 and over 1 orders of magnitude to maximum values of less than 0.5%. In 1 estuary, risk decreased by over 1 order of

magnitude, from amaximum of 36% to amaximum of 2.3%. At the other tributary and both the inner and outer estuaries, the

risk decreased to less than 1%, regardless of the tide. This analysis revealed that the remediation systemwas very effective and

that the standard level of protection required for slightly to moderately affected ecosystems (95% of species) by the Australian

and New Zealand Guidelines for Fresh and Marine Water Quality was met postremediation. Integr Environ Assess Manag

2012;8:98–106. � 2011 SETAC

Keywords: Ecological risk assessment Contaminated groundwater Remediation

INTRODUCTIONPenrhyn Estuary is a small (10 ha), tidal embayment

located approximately 10 km south of the Sydney centralbusiness district on the northern shoreline of Botany Bay,New South Wales (NSW), Australia (Figure 1). Land use inthe 320 ha catchment includes residential, commercial, andboth light and heavy industrial uses. The intertidal embay-ment is inundated at high tide, whereas at low tide, mudflatsare exposed. The estuary was originally devoid of vegetationwhen it was formed in the late 1970 s using sandy dredge spoilfrom development of the adjacent port. Today, however, itsupports a variety of flora species, including mangroves,saltmarsh species, and dune vegetation and also attractswading shorebirds, which forage on the mudflats at low tide.The fauna and flora are typical of that found in south easternAustralian marine and estuarine environments.

Groundwater contaminated with a complex mixture of atleast 14 volatile chlorinated hydrocarbons (VCHs) (AGEEandWoodward-Clyde 1990) has been delivered to the estuaryvia 2 drains (Springvale and Floodvale) that intercept thewater table since at least 1990 (AGEE and Woodward-Clyde 1990). Over 20 years of investigations and monitoringhave been undertaken at the site and its surroundings.

Summaries of many of these reports are available online(www.oricabotanytransformation.com/). The VCHs in thegroundwater and estuary are characterized by high aqueoussolubility, low octanol-water partition coefficients (KOW)(log KOW less than 3), and thus they have a low potential tobioaccumulate (Carey et al. 1998). VCHs act under thenonpolar narcotic mode of action (McCarty and Mackay1993; Carey et al. 1998).

The potential environmental impact that the contaminatedgroundwater may pose to the Penrhyn Estuary ecosystemswas previously evaluated using a tiered ecological riskassessment framework. This consisted of a hazard assessment(Hunt et al. 2007), followed by a probabilistic ecological riskassessment (PERA) (Hunt et al. 2010) that used the JPCmethod. The Hunt et al. (2007, 2010) assessments dividedthe estuary into 4 areas: Source Area (the Springvale andFloodvale drains), tributaries (the Springvale Tributary [SVT]and the Floodvale Tributary [FVT]), the inner estuary (IE),and the outer estuary (OE) (Figure 1). As the tributaries andestuaries are tidal, the risk was also assessed on a temporalbasis (at low, high, and mid tide). Spatially, the risk wasidentified as being greatest in the Source Areas with risk of upto 84% (that is equivalent to 80% of species being affected75% of the time). Risk further decreased in the followingorder: SVT (up to 36%), the IE (up to 9.5%), the FVT (up to2.6%), and was lowest for the OE (up to 0.8%). The ERAcalculated risk as being greatest at low tide followed by midtide then high tide. The key process reducing concentrationsof VCHs in the estuary was identified as dilution withseawater entering from Botany Bay (URS 2005).

Integrated Environmental Assessment and Management — Volume 8, Number 1—pp. 98–10698 � 2011 SETAC

* To whom correspondence may be addressed:

[email protected]

Published online 9 May 2011 in Wiley Online Library

(wileyonlinelibrary.com)

DOI: 10.1002/ieam.220

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In 2005, a groundwater treatment plant (GTP) thatintercepted and decontaminated the groundwater was com-missioned (Stening et al. 2008). The GTP is a ‘‘pump andtreat’’ groundwater system that extracts groundwater from anetwork of 113 wells and has a capacity to extract and treatup to 15 ML d�1 of groundwater. The treatment processinvolves stripping VCHs from the groundwater followed bydestruction at high temperature using a thermal oxidationunit (Stening et al. 2008). As a result of the GTP operation,the discharge of groundwater contaminated with VCHs to theSpringvale and Floodvale Drains decreased, resulting in areduced load of VCHs to Penrhyn Estuary (URS 2008). TheVCH load is highly variable over time with periods of highrainfall raising the groundwater elevation and subsequentdischarge regime, resulting in increased discharge to the drainsand transport to the estuary. The GTP achieved its objectiveof ‘‘hydraulic containment’’ of the groundwater (Stening et al.2008). Success of the remediation project has, however, onlybeen measured in terms of engineering (successful construc-tion and operation) and chemical (lower concentrations ofVCHs) criteria. Measurement of chemical concentrationsalone does not provide quantitative information on ecologicalrisk, merely qualitative information that risk should decreaseor increase.

Although risk assessments are commonly undertaken todetermine the current or predicted ecological risk associatedwith contamination, they are often not revisited afterremediation or the introduction of management practices.This occurs despite monitoring and feedback of changes inrisk being clearly included in various ERA frameworks (NEPC1999; USEPA 1998), being identified as vital to the success ofany risk assessment, and being central to the risk assessmentframework (Suter 1993; NEPC 1999). The value of riskassessment becomes limited if conditions change and the riskassessment is not updated (Burgman 2005).

The risk assessment of Penrhyn Estuary has not beenrevisited to evaluate whether the GTP has decreased theecological risk to organisms and ecosystems of the estuary.The objective of the present study was, therefore, to revisitthe risk assessment for VCH contamination of Penrhynestuary to assess the changed conditions and quantify thereduction in ecological risk after implementation of thegroundwater remediation program, and thereby, evaluatethe success of the project.

METHODS

Exposure assessment

To characterize exposure to VCHs in surface water in theestuary and source area, 9 sites were selected with 1 sitelocated in both Springvale and Floodvale drains (sites 1 and 2,respectively), 2 sites located in SVT (sites 3 and 4); 1 site inthe FVT (site 5); 3 sites in the IE (sites 6, 7, and 9); and 1 sitein the OE (site 8) (Figure 1). Sample sites were the same forboth the between preremediation and postremediationmonitoring programs. Springvale and Floodvale Drains inthe source areas have not been assessed for ecological risk perse, as these drains do not constitute ecosystems, but wereincluded for comparative purposes and source character-ization.

To characterize exposure to concentrations of VCHs insurface water in the estuary before the remediation, samplingof estuarine water and analysis for VCHs was undertaken overa 1-y period (in 2004 and 2005) in 2 monitoring programs asdetailed in Hunt et al. (2007, 2010). Samples were collectedat high and low tides and the programs aimed to characterizeshort- and long-term variability in VCH concentrations. Datawere compiled from all sites at high and low tides and used toquantify VCH concentrations at 3 exposure scenarios: 1) atmid tide (estimated as the average of the high and low tide

Figure 1. Sample sites in Penrhyn Estuary, Sydney, Australia. Dashed lines indicate the extent of each zone of the estuary and source areas.

Quantifying Reduction in Ecological Risk After Remediation— Integr Environ Assess Manag 8, 2012 99

concentrations), 2) high tide only, and 3) low tide only. The 3exposure scenarios were determined for the sites in thetributaries and inner and outer estuaries. Only 1 exposurescenario was undertaken for the source area sites as they arenot tidal.

Previous work (Hunt et al. 2007) has shown VCHconcentrations to vary temporally, therefore, to characterizeconcentrations of VCHs after the commencement of oper-ation of the GTP, 4 rounds of surface water surface watersamples (1 every 3 months) were collected between June2007 and March 2008. Data were compiled for the same 3exposure scenarios and the same 4 zones as the preremedia-tion ERA (Hunt et al. 2007, 2010). Five samples wereavailable from each of the 4 areas for assessment at low andhigh tides, with 12 samples available for the mid tides.

Samples were collected in 40mL glass vials with airtightTeflonTM lined lids with zero headspace, preserved withhydrochloric acid, and immediately stored at less than 4 8C.Samples were extracted using purge and trap methodology(US Environmental Protection Agency [USEPA] 5030B)and analyzed by gas chromatography mass spectrometry(GC/MS) using a modification of the USEPA Method8260B for volatile organic compounds (USEPA 1996) asdescribed in Hunt et al. (2009a). The limit of reporting was1mg/L for all analytes with the exception of vinyl chloride(10mg/L). Quality control evaluations were undertaken oneach of the sample batches and no analytes were detected inthe method blanks. Recoveries for laboratory control samplesand matrix spikes were between 80% and 120%, and withinthe acceptable criteria. Differences between primary andduplicate samples were generally less than 25%, typical of thevariability observed between duplicate samples for thesecontaminants at this laboratory and considered acceptable(Hunt et al. 2007, 2010; URS 2008). A sensitivity analysisundertaken in Hunt et al. (2010) concluded where resultswere less than the limit of reporting (LOR) there was littledifference in the outcome of the ERA when using either0.5 �LOR or 1 �LOR. However, substitution with 0 �LORresulted in difficulty in plotting the normal distribution.Consistent with the approach adopted in the previous PERAundertaken on Penrhyn Estuary (Hunt et al. 2010), valuesthat were less than the LOR were assigned a concentrationequal to half the LOR, considered a conservative approach(Warren Hicks et al. 2002). The distributions of each of theexposure data sets were assessed for log-normality using theAnderson-Darling test.

Effects assessment

Data for the effects assessment were developed using thedirect toxicity assessment (DTA) approach. The same specieswere used for the DTA as in the preremediation ERA (Huntet al. 2010) and details of the test methods are providedtherein. This is desirable because, providing there is no changein the sensitivity of the test organisms (that was checked withreference toxicant tests) any changes in the toxicity and riskbetween pre- and postremediation can be attributed tochanges in the environmental concentrations of VCHs, ratherthan due to using different species.

In the DTA, 2 groundwater samples (1 of shallowgroundwater discharge and 1 of groundwater from a nearbypiezometer, both upgradient of the estuary) were collectedfrom the source area and adjusted to marine salinity using

artificial salts. The 2 samples were combined in a ratio of9:1 to produce concentrations of VCHs high enough to elicita response in all of the test species. During testing,temperature, pH, salinity, and dissolved O2 content of arepresentative sample from each treatment were measureddaily and were within the required limits. Testing wasundertaken at a pH of between 7.6 and 8.3, at 20� 1 8Cand a salinity of 35� 1%. Reference toxicants and controls forartificial salt water and filtered seawater were undertaken foreach test.

Direct toxicity assessment of the contaminated ground-water was undertaken on 5 species: a 72-h algal (Nitzschiaclosterium) population growth test, a 72-h sea urchin(Heliocidaris tuberculata) larval development test, a 72-hoyster (Saccostrea glomerata, formerly Saccostrea commercialis)larval development test, a 96-h amphipod (Allorchestescompressa) survival test, and a 96-h juvenile polychaete(Diopatra dentata) survival test, that belong to 4 taxonomicgroups and thus meet the minimum data requirements forusing species sensitivity distribution (SSD) (ANZECC andARMCANZ 2000; Warne 2001). In accordance with therequirements outlined in ANZECC and ARMCANZ (2000),the test species were considered representative of thereceiving environment, ecologically relevant, and have somecommercial or recreational value in the area. To prevent lossof volatile contaminants and potential underestimation oftoxicity, toxicity testing was undertaken in sealed containersusing the methodology in Hunt et al. (2009a).

The toxicity data generated by the DTA is based on acute(A. compressa and D. dentata), subchronic (H. tuberculata andS. glomerata), and chronic exposure (N. closterium). The 2larval development tests for the urchin and oyster (subchronictests), were treated as chronic tests along with the algal testfor the purposes of guideline derivation and no acute tochronic ratio (ACR) was applied. The larval developmenttests were treated as chronic tests in the derivation of theSSD, because early life-stage (ELS) testing such as this can beconsidered chronic in the derivation of water quality guide-lines (Warne 2008; USEPA 2002). Further discussion of thesetests and their treatment as either chronic or acute is providedin Hunt et al. (2009b). To only use chronic toxicity data inthe SSD an ACR of 5 was applied to no observed effectconcentrations (NOEC) from the acute polychaete andamphipod tests, in accordance with the published ACRvalues of 4.5� 2.5 (McGrath et al. 2004) and 5.09� 0.95 (DiToro et al. 2000). The effect of ACR size and treatment ofsubchronic tests as chronic tests was evaluated in Hunt et al.(2009b)

Hunt et al. (2009b) compared the fit of the reciprocalPareto, (the distribution determined by the software used toderive the Australian and New Zealand water quality guide-lines, i.e., BurrliOZTM) (Campbell et al. 2000), and log-normal distributions for the preremediation DTA data andconcluded that the log-normal gave the best fit to the data.The log-normal distribution was therefore adopted in thecurrent study.

Risk characterization

Hunt et al. (2010) and others (Aldenberg et al. 2002; vanStraalen 2002) have identified the multiple benefits of usingthe JPC methodology to characterize risk, including thequantification of the risk (d, the area under the curve) and

100 Integr Environ Assess Manag 8, 2012—J Hunt et al.

provision of information on the type of exposure (the shapeof the curve). When using the JPC to interpret ecological risk,the greater the area under the curve (AUC), the greater theecological risk, whereas the shape of the curve allowsdifferentiation between a high damage/low probability eventand a low damage/high probability event. Risk values (d) andJPCs were estimated using the ETXTM program (vanVlaardingan et al. 2004), which estimates log-normal curvesfor each of the exposure and toxicity distributions and theextent of overlap between these 2 distributions. Risk wascharacterized for each of the 3 exposures scenarios (data atmid, high, and low tides) using NOEC toxicity data for sitesin the tributaries and the inner and outer estuaries, resultingin a total of 12 risk values. Risk was characterized for a singleexposure scenario using NOEC toxicity data for sites in thesource area, resulting in 2 additional risk values. Wherestandard deviations of exposure data were too large for ETXto calculate an area under the curve (AUC), these wereestimated manually. Both the previous (Hunt et al. 2010) andcurrent PERA used a threshold of 5% as the acceptable risk,which is similar but not identical to the 95% protectionapplied in the Australian and New Zealand Guidelines forFresh and Marine Water Quality (ANZECC and ARM-CANZ, 2000). The threshold of 5% risk in the previous ERAis the concentration at which 5% of species would always beaffected (100% of the time), whereas the PC95 in ANZECCand ARMCANZ is the concentration at which 5% of specieswould be affected 50% of the time.

RESULTS AND DISCUSSION

Exposure assessment

Mean concentrations of VCHs in the Springvale Drainsource area decreased by approximately 2 orders of magni-tude, from 22036 to 250mg/L (Table 1), following commis-sioning of the GTP. In Floodvale Drain source area, the meanconcentration of VCHs decreased by 1 order of magnitude,from 1420 to 110mg/L after remediation. In the tributariesand the estuary, mean concentrations postremediation werebetween 5-fold (in the SVT at high tide) and approximately40-fold (in the IE at low tide) lower in all locations andexposure scenarios.

Postremediation exposure concentrations generally fol-lowed a similar trend to preremediation exposure concen-trations, i.e., concentrations were generally greatest at lowtide followed by mid tide and high tide. When assessedspatially, preremediation exposure concentrations in non-source areas were generally greatest in SVT followed by FVTand the IE, with the lowest concentrations in the OE.Postremediation, the trend was the same except that the FVThad concentrations lower than IE (Table 1). Concentrationsof VCHs exhibited a high degree of spatial and temporalvariability postremediation, similar to preremediation expo-sure concentrations.

When log-normality of postremediation exposure distribu-tions was assessed using the Anderson-Darling test, 50% ofexposure scenarios failed (at p< 0.05). Scenarios that failedwere those in the IE and OE, where large number of samplesrecorded concentrations of VCHs that were less than the limitof reporting. In the previous ERA, approximately 25% ofexposure scenarios were not log-normal (Hunt et al. 2010).The scenarios where the test for log-normality failed wereexamined and the exposure (mean of 11mg/L; maximum of22mg/L) concentrations were considerably less than the site-specific PC95 of 830mg/L, discussed below. The risk wastherefore considered to be acceptably low at these locationsand the lack of log-normality of the distribution was notconsidered to affect the assessment of risk.

Effects assessment

Postremediation toxicity metrics including NOEC, lowestobserved effect concentrations (LOEC), and effect concen-trations to 50% of organism (EC50) were derived for theDTA tests in Hunt et al. (2009b) and expressed in terms ofconcentrations of total VCHs (Table 2). NOECs varied from1.11mg/L for urchin larval development test to 45.5mg/L forthe amphipod survival test. Similar variations in toxicityoccurred for the LOEC and EC50 data (Table 2). The toxicitydata were log-normally distributed when assessed using theAnderson-Darling test (p< 0.05). A log-normal SSD wasderived and the concentration that should theoreticallyprotect 95% of species from experiencing toxic effects(PC95) was 830mg/L. The PC95 is equivalent to the

Table 1. Concentrations of volatile chlorinated hydrocarbons in surface water in Penrhyn Estuary, Australia

RemediationTidalstatus

Springvaledrain (mg/L)

Floodvaledrain (mg/L)

SVT(mg/L)

FVT(mg/L)

IE(mg/L)

OE(mg/L)

Mean SD Mean SD Mean SD Mean SD Mean SD Mean SD

Before No tide 22036 18865 1420 685 — — — — — — — —

Mid — — — — 1816 2014 329 248 298 671 51.2 92.3

High — — — — 1363 2436 132 146 91.5 83.3 31.8 42.6

Low — — — — 2273 2098 419 141 669 1013 151 161

After No tide 250 200 110 191 — — — — — — — —

Mid — — — — 187 152 10.8 5.7 13.4 15.9 9.0 7.5

High — — — — 161 113 9.7 3.7 9.1 7.6 9.8 9.6

Low — — — — 214 194 11.9 7.6 17.7 21.3 8.3 5.5

FVT¼ Floodville Tributary; IE¼ inner estuary; OE¼outer estuary; SD¼ standard deviation; SVT¼ Springvale Tributary.


concentration that should permit only 5% of species toexperience toxic effects (the HC5) more commonly used inEurope.

Risk characterization

Using the JPC approach, AUCs were derived for each ofthe 3 exposure scenarios (mid tide concentrations, high tideconcentrations, and low tide concentrations), for each of the 4estuary areas (SVT, FVT, IE, and OE), and a single exposurescenario for the 2 sources resulting in a total of 14 risk (d)values for pre- and postremediation (Table 3). Environmentalrisk, measured as d, was lower in all locations and all exposurescenarios after commissioning of the GTP, except for the hightide scenario in the OE, where the risk did not change(d remained at 0).

Overall, although the magnitude of environmental riskdecreased in the estuary, the spatial trend in risk postreme-diation remains similar to that preremediation, which reflectsthe underlying physical characteristics of the source areas andtheir interaction with Penrhyn Estuary and Botany Bay.Before remediation, the risk in the estuary decreased from thegreatest risk in SVT >> FVT � IE>OE; however, post-remediation, the risk was greatest in SVT >> FVT¼ IE¼OE.

Before the remediation, Hunt et al. (2010) hypothesized thatthe greater risk in the IE reflected the greater input ofcontaminants from the Springvale Drain source area. Aftergroundwater remediation, a greater decrease in risk wasidentified in the Springvale Drain source area than theFloodvale Drain source area (Table 3). The greater decreasein risk in the IE than in FVT due to the remediation mayreflect the greater influence of the SVT on the magnitude ofcontamination in the estuary, than the FVT.

In the Springvale Drain source area, risk decreased fromapproximately 84% preremediation, to less than 0.5%postremediation. In Floodvale Drain source area, riskdecreased from approximately 16% to less than 0.5%(Table 3). In the estuary, after remediation, the mean riskvalue (d) across the 4 locations and both tides, decreased from8% to less than 1%. Decreases in the magnitude of risks ineach location demonstrate the strong positive impact of thegroundwater treatment system on concentrations of VCHsdischarging to Penrhyn Estuary. The ecological implication ofthe remediation can be assessed directly from the JPCs. Itshould be noted that the JPCs are based on site-specific datafor only 5 species, and although this is the minimumacceptable for such site-specific assessments in Australia(ANZECC and ARMCANZ, 2000), it is acknowledged that

Table 2. Summary of NOEC, LOEC, and EC50 values for toxicity testing ofmarine test organisms exposed to groundwater contaminated withvolatile chlorinated hydrocarbons

Toxicity metricAlga(mg/L)

Urchin(mg/L)

Oyster(mg/L)

Polychaete(mg/L)

Amphipod(mg/L)

NOEC 2.30 1.11 4.98 29.9 45.5

LOEC 4.98 2.30 10.3 45.5 >45.5

EC50 4.10 3.77 9.79 32.1 >45.5

EC50 95% LCL 2.32 3.57 9.23 27.16 —

EC50 95% UCL 7.13 3.98 10.32 38.05 —

EC50¼ effective concentration, 50%; LCL¼ lower confidence limit; LOEC¼ lowest observed effect concentration; NOEC¼no observed effect concentration;

UCL¼upper confidence limit.

Acute-to-chronic ratios have not been applied. Data were originally presented in Hunt et al. (2009b).

Table 3. Risk values (%) before and after implementation of the groundwater treatment plant based on preremediation DTA results

Implementation ofthe treatment plant

Tidalstatus

Springvaledrain

Floodvaledrain SVT FVT IE OE

Before No Tide 83.8 15.9 — — — —

Mid — — 25.0 2.6 4.1 0.1

High — — 16.0 0.1 0.1 0.0

Low — — 35.4 2.3 9.3 0.8

After No Tide 0.37 0.43 — — — —

Mid — — 0.91 0.0 0.0 0.0

High — — 0.31 0.0 0.0 0.0

Low — — 2.26 0.0 0.0 0.0

Dash indicates no risk value was derived, because these locations are not tidal; DTA¼direct toxicity assessment; FVT¼ Floodville Tributary; IE¼ inner estuary;

OE¼outer estuary; SVT¼ Springvale Tributary.


the error in PC estimates increases as the number of datapoints decreases (Pedersen et al. 1994; Newman et al. 2000;Wheeler et al. 2002). Therefore, although the valuesgenerated by the JPC are discussed in the following, theyshould be not be taken to be absolute values but ratherindicative of the level of protection that is likely to beprovided.

In the Springvale Drain source area, preremediation, lessthan 5% of species would be protected 50% of the time(Figure 2a), however, postremediation greater than 95% ofspecies would be protected at least 95% of the time(Figure 2b). In the SVT preremediation at least 90% ofspecies would be protected approximately 50% of the time(Figure 2c), however, postremediation greater than 95% of

Figure 2. Joint probability curves for ecological risk for preremediation and postremediation scenarios in the Springvale drain source area (a,b), the Springvale

Tributary (c,d), and the inner estuary (e,f) across both tides.


species would be protected at least 95% of the time(Figure 2d). In the IE, the risk profile indicated a changefrom 90% of species being protected 90% of the timepreremediation (Figure 2e) to 100% of species beingprotected 100% of the time (Figure 2f) postremediation.

Before remediation, the greatest mean ecological risk wasat low tide (mean d¼ 12, n¼ 12), followed by mid tides(mean d¼ 8.0, n¼ 12), with the lowest risk reported at hightide (mean d¼ 4.0, n¼ 12) (Hunt et al. 2010). Afterremediation, the greatest mean ecological risk was at lowtide (mean d¼ 0.34, n¼ 4), followed by mid tides (meand¼ 0.21, n¼ 4), with the lowest risk identified at high tide(mean d¼ 0.20, n¼ 4). These interpretations of risk areconsistent with the physical characteristics of the estuary,where concentrations of VCHs and risk are highest at low tideand lowest at high tide and reflect the overall pattern reportedin the screening level risk assessment, where risk at lowtide>mid tide>high tide (Hunt et al. 2007) and tidalinteraction between the OE and Botany Bay.

The risk values after remediation were considerably lessthan the 5% threshold level of acceptability for risk adopted inthe present study and Australian water bodies in general,indicating that the GTP has successfully reduced themagnitude of ecological risk in the estuary to an acceptablelevel. Together with the risk values, the JPCs also reflect thestrong positive impact of the GTP on VCH contamination inthe estuary and suggest that VCH contamination hasdecreased to an acceptably low level and recovery of theaquatic ecosystems would be expected. Measurement ofconcentrations of VCHs alone did not provide quantitativeestimates of risk, merely quantitative information on expo-sure. Reassessment of the risk assessment is vital to thesuccess of the ERA process and a central feature of the riskassessment framework.

The chemical composition has changed since the prereme-diation ecological risk assessment was undertaken. Prereme-diation surface water from Penryhn Estuary was dominatedby dichloroethane, however, since the remediation, thecomposition has been increasingly dominated by dichloroe-thene and vinyl chloride. These compounds are bothdegradation products of higher order VCHs, however, vinylchloride, in particular, poses problems for undertakingtoxicity testing as it is a gas at standard temperature andpressure (and highly carcinogenic). The preremediation DTA(Hunt et al. 2010) was undertaken at ranges of up to 50mg/Ltotal VCHs and toxic responses did not result for some testspecies. Given the low VCH concentrations present at the sitepostremediation (total VCHs of <1mg/L), it was consideredvery likely that a new round of DTA would result in even lessfrequent responses and smaller magnitude responses. Thiswould make it difficult to undertake a risk assessment usingthe probabilistic methods outlined in the previous and currentrisk assessments. The following text explains how weattempted to overcome this limitation and the results of thesubsequent ERA.

The total internal lethal residue values (ILC50) of theVCH mixture, chloroform, and dichloroethane for the 5 testorganisms were 0.9, 2.8, and 2.3mmol/kg, respectively.These results support the findings of McCarty and Mackay(1993) and Carey et al. (1998) that VCHs all have a non-polar narcotic mode of action (ILC50 should range from 1 to10mmol/kg). From a theoretical perspective (Plackett andHewlet 1952), mixtures that contain chemicals with the same

mode of action and that do not interact should exert a toxicitythat conforms to the concentration addition (CA) model ofjoint action. This theoretical prediction has repeatedly beenshown to be valid (refer to literature cited in Dyer et al.2011). In addition, a number of studies have shown that theCA model provides a more conservative estimate of mixturetoxicity than the independent action model of mixturetoxicity (Faust et al. 1994; Backhaus et al. 2000a, 2000b;Dyer et al. 2000; Junghans et al. 2006; Chevre et al. 2006)that applies where mixtures contain chemicals that do notinteract but have different modes of action. Given thepreceding, frameworks for estimating the toxicity of mixtureshave been developed (Altenberger et al. 2004; Junghans2004; De Zwart and Posthuma 2005; Posthuma et al. 2008;Dyer et al. 2011) all of which recommend the use of the CAmodel in the first instance. Therefore, the CA model was usedto provide an estimate of the risk posed by the mixture ofVCHs in the surface waters of Penrhyn Estuary. The toxicunits necessary for the CA calculation were calculated bydividing the concentrations of individual contaminants bytheir corresponding (low reliability) trigger values from theAustralian and New Zealand Guidelines for Fresh and MarineWater Quality (ANZECC and ARMCANZ 2000) and thosederived by Hunt et al. (2007). These TU values for eachcomponent were then summed for each of the 4 samplescollected at each site at low, mid, and high tides. The originaltoxicity curve (presented as total VCHs in Hunt et al.[2009b]) and the exposure curves were recalculated based ontoxic units and the CA model. The exposure curves werederived for each zone of the estuary (2 source zones, 2tributaries, an IE and OE) at the 3 tide scenarios for all 4samples collected during the years 2007–2008. This ensuresboth the spatial and temporal variability of the VCHcomposition and concentrations are considered. It should benoted that in this ERA, however, that both types of variabilityare only considered for 2007–2008 due to the time frame ofthe sampling. Subsequent changes or variability can not beconsidered without further sampling.

Using the JPC method described above, the risk wasrecalculated. This provides a mechanism for undertaking aprobabilistic risk assessment using the available toxicity data(derived for the site) and the contaminant mixture from themost recent rounds of monitoring and the incorporates thechanges in composition of the mixture. The risk for themixtures of VCHs measured during a year (June 2007–March2008) indicated that an unacceptable risk (>5%) remains in

Table 4. Quotient-based risk values (%) after implementation of thegroundwater treatment plant calculated using the concentration

addition model for mixture toxicity

Tidal statusSpringvale

drainFloodvale

drain SVT FVT IE OE

No Tide 7.76 0.62 — — — —

Mid Tide — — 10.32 0.09 0.18 0.17

High Tide — — 3.52 0.11 0.13 0.28

Low Tide — — 15.52 0.07 0.28 0.1

FVT¼ Floodville Tributary; IE¼ inner estuary; OE¼outer estuary; SVT¼ Spring-

Springvale Tributary.


the Springvale Source Area (7.8%; Table 4) and in the SVT atlow tide (15%) and mid tide (10%). For all other combina-tions of site and tide in the estuary, the risk was less than 0.3%and therefore deemed acceptable (ANZECC and ARMCANZ2000). It should be noted, however, that the toxic units forthis assessment were based on low reliability trigger values.These are derived using a minimal data set and by dividingthe lowest toxicity value by an assessment factor of 1000.These TVs are likely to be smaller than if a larger morerepresentative toxicity data set was available. Therefore, theERA determined above is likely to overestimate the actualrisk posed.

Uncertainty

The purpose of including a discussion of uncertainty in therisk assessment is to inform risk managers and decision makersof the uncertainty that exists with the information presented.As was the case for the preremediation PERA, the treatmentof uncertainty presented here only identifies sources ofuncertainty and does not convey the potential extent orimpact of the uncertainty on the risk assessment. It isimportant, nonetheless, to be explicit with all the sources ofuncertainty to ensure that the risk assessment is transparent(Calow 1998). Sources of uncertainty, discussed in detail inHunt et al. (2010), included toxicity testing undertaken atconstant exposure concentrations (pH and salinity), the use ofSSDs with the inherent assumption that protection of aproportion of species will protect ecosystem structure andfunction, the use of NOEC data in the derivation of SSDdistributions, the application of log-normal distributions toright-skewed exposure data, and changes in the chemicalcomposition of the discharge to the estuary.

Penrhyn Estuary, being intertidal, contains areas ofvarying pH and salinity. Therefore, there is some uncertaintyin evaluating the risk in an environment of varying pH andsalinity using the results of toxicity testing undertaken atstatic pH and salinity. However, the toxicity of the VCHs inthis study is not known to be influenced by changes in pH orsalinity. The authors considered this when undertaking thebioassays, however, the decision was made to undertakemarine testing (in marine conditions) as the majority of theestuary (inner, outer, and lower reaches of both tributaries)are characterized by marine conditions. Testing was main-tained at marine conditions to reduce potential confoundingand ultimately, to represent the receiving ecosystem, the OEand Botany Bay.

Many authors have criticized the use of NOEC data(Chapman et al. 1996; Newman 2008; Warne and Van Dam2008) for a variety of reasons but principally because of theinappropriateness of the statistical methods used in theirderivation and that the NOEC can only be one of theconcentrations used in toxicity tests. Despite these criticisms,NOEC data continue to be generated and used, largelybecause no authoritative organization has banned their useand they are incorporated into regulations and laws.Regulatory authorities have tried to limit the effect of theNOECs only being a tested concentration by specifying amaximum difference between treatments (e.g., the OECDrecommends a maximum factor of 3.2). In the current projectthe difference between test treatments was 2-fold, thereforethe maximum difference that can occur between the true noeffect concentration and the NOEC is a factor of 2.

CONCLUSIONSThe present study quantified the reduction in ecological

risk posed by VCH contamination in Penrhyn Estuary aftercommissioning of a groundwater treatment plant. The site-specific nature of the toxicity and exposure distributionsgreatly increased the relevance of the risk assessment.

The ecological risk assessment, based on direct toxicitytesting undertaken before the remediation, shows that thegroundwater remediation has had a strong positive impact onthe conditions in the estuary and that the risk from VCHcontamination has decreased to acceptably low levels, whererecovery of the aquatic ecosystem would be expected.

Some changes in the chemical composition of the ground-water were identified, in particular an increase in degradationproducts. Additional probabilistic assessment of the risk usinga toxicity quotient approach indicates the risk is acceptablewith the exception of 1 source area (at all times) and itstributary at low tide and mid tides.

Ongoing chemical monitoring should be undertaken ifchemical concentrations do not continue to decrease declineor the composition does not continue to change to less toxicVCHs, further DTA should be considered. Given the alreadylow VCH concentrations present, the VCH contaminantswould need to be concentrated before toxicity testing couldpermit the quantification of ecological risk.

Acknowledgment—This work was funded by OricaAustralia, with additional support from ALS Environmentaland Ecotox Services Australasia.

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