Science of the Total Environment 409 (2011) 2029–2039
Contents lists available at ScienceDirect
Science of the Total Environment
j ourna l homepage: www.e lsev ie r.com/ locate /sc i totenv
Organochlorines and heavy metals in wild caught food as a potential human healthrisk to the indigenous Māori population of South Canterbury, New Zealand
Michael Stewart a,⁎, Ngaire R. Phillips a, Greg Olsen a, Christopher W. Hickey a, Gail Tipa b
a National Institute of Water and Atmospheric Research, PO Box 11 115, Hamilton 3251, New Zealandb Tipa & Associates, 44 Chain Hills Road, RD1, Dunedin, New Zealand
Increasing concentrations of anthropogenic contaminants in wild kai (food) of cultural, recreational andeconomic importance to the indigenous Māori of New Zealand is a potential human health risk. Contaminantsthat are known to bioaccumulate through the food chain (e.g., organochlorine pesticides (OCPs), PCBs andselected heavymetals) were analysed in important kai species including eel (Anguilla sp.), brown trout (Salmotrutta), black flounder (Rhombosolea retiaria) and watercress (Nasturtium officinale) from importantharvesting sites in the region of South Canterbury. Eels contained relatively high wet weight concentrationsof p,p′-DDE (8.6–287 ng/g), PCBs (32ΣPCB; 0.53–58.3 ng/g), dieldrin (b0.05–16.3 ng/g) andΣchlordanes (0.03–10.6 ng/g). Trout andflounder contained lower concentrations of organochlorines than eels,with p,p′-DDEwetweight concentrations ranging from 2.2 to 18.5 ng/g for trout and 6.4 to 27.8 ng/g for flounder. Total arsenicwet weight concentrationswere below detection limits for eels but ranged from 0.27 to 0.89 μg/g for trout and0.12 to 0.56 μg/g for flounder. Mercury concentrations ranged from0.02 to 0.56 μg/g, 0.11 to 0.50 μg/g and 0.04to 0.10 μg/g (ww) for eel, trout and flounder respectively. Lifetime excess cancer risk was calculated throughestablished risk assessment procedures, highlighting dieldrin, ΣPCBs and p,p′-DDE in eels and arsenic in troutand flounder as primary contaminants of concern. A second non-cancer chronic health risk assessmentindicated that mercury and PCBs were a potential concern in eels and mercury in trout. A cumulative lifetimecancer risk assessment showed potential health risk for consumption of some species, even at lowconsumption rates and provided the basis for establishing recommended dietary consumption limits forharvest sites within the study region.
Traditionally, indigenous New Zealand Māori had their ownknowledge systems of how the environment contributed to healthandwell-being.Wild kai (food), gathered from the sea, rivers, and lakes,has always been of significant cultural, recreational and economicimportance in both traditional and contemporary Māori societies.Historically, traditional Māori knowledge recognised many freshwaterspecies in the fauna, andMāori valued several of themas significant foodresources. This was especially true of the ubiquitous, large and highlynutritious eels (Anguilla sp.), but also the lamprey (Geotria australis),whitebait and kokopu (Galaxias spp.) and others (McDowall, 2000). TheNew Zealand longfin eel (Anguilla dieffenbachii) reaches a length ofabout 2000 mmandweighs up to 25 kg, while the shortfin eel (Anguillaaustralis) reaches 1300 mm and may weigh 3.5 kg. The lamprey
(G. australis) grows to 900 mm, and the giant kokopu (Galaxiasargenteus) has been recorded up to 580 mm, with a weight of nearly3 kg (McDowall, 2000). In some areas, a range of well developed traps,nets and devices were used to ensure a large harvest of fish andcrustaceans (Hiroa, 1921). Substantial quantities of watercress(Nasturtium officinale), first introduced to New Zealand freshwatersabout 1850, were also harvested. However, as watercress stronglybioaccumulates metals, and is a hyperaccumulator of arsenic – due toincorporation into iron oxides attached to the plant surface (Robinsonet al., 2003) – it represents a potential human health risk incontaminated environments. Marine kai were also important andincluded a diverse range of abundant shellfish, crustaceans and fishspecies. Harvesting of wild caught kai, particularly in freshwater, hasdeclined over time and has been replaced predominantly by a diet ofstore bought food, similar to that of the average non-Māori NewZealander (Tipa et al., 2010a, 2010b).
A recent review of wild food in New Zealand (Turner et al., 2005)identified gaps in the knowledge of contaminants in non-commercialwild-caught foods, especially in terms of consumption levels (andhence exposure). A resulting draft position paper (NZFSA, 2005)
2030 M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
identified the need for information and education on contaminants inkai. In addition, while existing consumptive advice is available forsome species of relevance to Māori, this advice is based on averagenational consumptive patterns and doesn't account for potentiallyhigher consumption rates of specific, traditionally harvested foods byMāori, with its concomitant elevated exposure risk.
The majority of the international research in the area ofcontaminants in the traditional diets of indigenous peoples hasprimarily focused on consumption levels and health effects ofexposure to heavy metals and organochlorine contaminants throughthe consumption of marine fish and mammals in people from thenorthern hemisphere. For example, international indigenous contam-inant research programmes such as the Northern ContaminantsProgramme (NCP) and the Effects on Aboriginals from the GreatLakes Environment (EAGLE) Project were established in response toconcerns regarding the exposure of humans to elevated levels ofcontaminants in the traditional subsistence diets of indigenouspeoples. Research to date has shown that certain indigenouscommunities have elevated contaminant levels due to exposurethrough their traditional diet (Hoekstra et al., 2005; Johansen et al.,2004; Odland et al., 2003; Van Oostdam et al., 1999, 2003).
As many toxic contaminants are stored in the lipids of biota theycan be biomagnified up the food-chain. For example, eels are lipidrich, are top of the food chain in freshwater aquatic environments inNew Zealand and are therefore likely to contain the highestconcentrations of bioaccumulative contaminants. It is unlikely thatcontemporary Māori communities – with their supplementary wildkai diets – have been exposed through their diet of wild kai to levels ofbioaccumulative contaminants as high as those observed in indige-nous populations residing in the northern hemisphere, particularlybecause of the absence of top level mammalian predators in the Māoridiet. However, risks to Māori of wild kai dietary consumption has notbeen quantified.
Bioaccumulative contaminants that are of potential concern are theorganochlorine pesticides (OCPs, i.e., DDTs, dieldrin and lindane),polychlorinated biphenyls (PCBs), pentachlorophenol and dioxins,polycyclic aromatic hydrocarbons (PAHs), as well as certain heavymetals such as mercury, arsenic, cadmium, lead, copper and zinc. NewZealandused a considerable amount of organochlorine pesticides fromthe 1940s to the 1970s. DDT, in particular, was used largely to controlgrass grubs and porina caterpillars, with its use restricted in 1970 andfinally banned in 1989 (Taylor et al., 1997). Canterbury is a regionwitha large agriculture and horticulture industry, where the application oforganochlorine pesticides was widespread. Although a nationwidesurvey on organochlorines, including PCBs, was carried out in 1995(Buckland et al., 1998a), the region of South Canterbury was excludedfrom this study. In addition, sheep dips were arsenic-based until the1950s, with organochlorine (e.g., dieldrin, lindane and DDT) andorganophosphate (e.g., diazinon) insecticides used after this time(Environment Canterbury, 2010). There are thought to be over 50,000contaminated sheep-dip sites in New Zealand (MfE, 2006). Metalssuch as mercury and arsenic can enter into the food chain due toelevated environmental levels from geothermal inputs and/or asseepage from urban refuse stations. Cadmium, lead, copper and zincare associated with urban contamination, usually as diffuse sources.Cadmium is also elevated in the widely used rock phosphate fertiliser,which has been attributed to high concentrations measured in marineshellfish in some areas (Butler and Timperley, 1996). The consumptiverisk for many of these contaminants is potentially and significantlyelevated with the consumption of long-lived high fat freshwaterspecies, such as eels, as compared with other fish diets.
In this paper, we describe the results of a survey of sitestraditionally associated with the gathering of wild kai by local Māori.We characterise the concentrations of bioaccumulative contaminantsin key commonly gathered animal and plant species, as well as inassociated aquatic sediment samples. We then undertook a risk
assessment based on local and New Zealand-wide consumption ratesfor key species. Finally, we discuss the implications of these results forMāori and non-Māori communities.
2. Materials and methods
2.1. Survey design
Information on kai harvesting (i.e., site and species) was collatedfrom the results of collective meetings with members of anindigenous Māori population belonging to Te Runanga o Arowhenua(tribal assembly of Arowhenua) located in the South Canterburyregion (Stewart et al., 2010). Analysis of this information allowed forthe design of a sampling regime which characterised contaminantconcentrations in food and the associated environment (sediment),from sites of direct relevance to participants. In addition, a survey ofpast and present consumption patterns was undertaken by question-naire with this same group and detailed individual interviews(n=12) of dietary intake, to establish historic and contemporaryconsumption rates of key species. Indicative consumption ratesranged from less than once per month to one or more times per day(Tipa et al., 2010a).
2.2. Sampling
The focus of this study was South Canterbury, New Zealand, anarea including the towns of Timaru (population 36,500), Temuka(pop. 4000) and Geraldine (pop. 2200) south of the Rangitata River(Fig. 1). In the wider Timaru district of 42,867 people, those whoidentify as Māori make up 6.1% of the population (Statistics NewZealand, 2010).
The major river networks include the Temuka (611 km2; medianflow 3.5 m3/s), Orari (714 km2; median flow 12.8 m3/s) and OpihiRivers (628 km2; median flow 20.3 m3/s) and Te NgaWai River inland(a tributary of the Opihi River; 488 km2; median flow 4.9 m3/s). TheTe Nga Wai River was considered the reference site for all othersampling sites and with the Temuka River forms part of the OpihiRiver catchment of about 2450 km2.
The headwaters of the Opihi River are found in the foothills of theSouthernAlps at elevationsup to 2200 m.Only 0.4% of the catchment isindigenous bush, with 30% tussock grassland. Extensivemixed grazingoccurs in the foothills, intensive pastoral farming on the lowlands andintensive land use (cropping, horticulture, and dairy farming pasturewith irrigation) on the plains (Environment Canterbury, 2000). Thelower reaches of the Opihi River have a braided river channel with acobble substrate.
The Opihi River receives point source wastewater discharges fromthree domestic sewage oxidation lagoons (Burkes Pass, Fairlie andPleasant Point), the Temuka River discharges from two oxidationlagoons (Geraldine and Temuka) and rinse waters from one woolscour plant (Environment Canterbury, 2000). Urban stormwaters andagricultural run-off comprise additional contaminant sources. Wide-spread historic use of organochlorine pesticides (mainly DDT anddieldrin) has resulted in areas of contaminated land, with potentialdiffuse inputs of these legacy contaminants.
The full information pertaining to collection of biota is contained inthe Supplementary information (Table S1). However, briefly, as aninitial screening exercise and invoking the desire of local iwi (Māoritribal group) to collect minimal numbers of some culturally importantspecies, we surveyed 12 sites and collected comparably sizedindividual specimens of selected species from each site (if present).All fish were caught by electric fishing techniques, with the exceptionof Opihi river mouth and Orari Ohapi, where nets were used to catchtrout. Watercress was harvested by hand, avoiding roots. A total of 9short fin eels (A. australis), 1 long fin eel (A. dieffenbachii), 5 browntrout (S. trutta) and 4 black flounder (R. retiaria) were collected. All
0 10 km
N
TIMARU
Washdyke Creek
Washdyke Lagoon
Doncaster
Opihi Below Pleasant PointOpihi River Mouth
Opihi Upstream
Te Nga Wai River
Orari OhapiTemuka TEMUKA
Te Nga WaiOhapi CreekOpihi River
Winchester
Waihi River
GERALDINE
Rangitata River
Orari River
0 10 km
N
TIMARU
Washdyke Creek
Washdyke Lagoon
Doncaster
Opihi Below Pleasant PointOpihi River Mouth
Opihi Upstream
Te Nga Wai River
Orari OhapiTemuka TEMUKA
Te Nga WaiOhapi CreekOpihi River
Winchester
Waihi River
GERALDINE
Rangitata River
Orari River
0 10 km0
NN
TIMARU
Washdyke Creek
Washdyke Lagoon
Doncaster
Opihi Below Pleasant PointOpihi River Mouth
Opihi Upstream
Te Nga Wai River
Orari OhapiTemuka TEMUKA
Te Nga WaiOhapi CreekOpihi River
Winchester
Waihi River
GERALDINE
Rangitata River
Orari River
Fig. 1. Collection sites in this study with map of South Island of New Zealand (inset) showing location of South Canterbury region.
2031M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
samples were kept on ice and then frozen until processing. Collectionswere undertaken in 2009, either between 12th and 14th May, or on3rd June. Composite watercress (N. officinale) samples were collectedfrom eight sites between 12th and 14th May. Composite surficialsediments (top 0–2 cm) were collected from all sites, at the time ofbiota collection, from areas where the finer particle size fraction waslikely to be deposited.
2.3. Sample preparation
Eels, trout and flounder were partially thawed, weighed andmeasured. Clean fillets of fish or eel muscle tissue were carefullyremoved from each individual, avoiding the gut. Otoliths wereremoved from 9 eels for accurate age determination. Watercress wascut, while frozen, into small pieces. All samples were weighed, frozenand freeze driedwith a shelf temperature of−20 °C. Biometric data foreel, trout/flounder and watercress are presented in Supplementaryinformation Tables S2–S4.
Each sediment composite was allowed to thaw and placed in ashallow plastic tray. All large stones and plant material were removedand the sediment thoroughly homogenised before freeze drying.Freeze dried sediment was sieved dry through a 2 mm stainless steelsieve and all material greater than 2 mm discarded. A sub-sample ofthe freeze dried (b2 mm) sieved sediment was re-suspended inNanopure water, sonicated for 1 h and wet sieved through a 63 μmnylon mesh. The b63 μm fraction was oven dried and a gravimetricanalysis performed. Sediment size analysis data are presented inSupplementary information Table S5.
2.4. Analysis of contaminants in kai and sediment
For OCPs and PCBs a procedure based on accelerated solventextraction (ASE), gel permeation chromatography, silica/alumina clean-up and gas chromatography–mass spectrometry (GC–MS) was used,closely following the published procedures of United States Environ-mental Protection Agency (US EPA, 1977, 1986) and National Oceanicand Atmospheric Administration (NOAA, 1993). Briefly, sedimentsamples (either 10 g or 20 g) or freeze dried fish tissue (6 g) werespiked with analytical surrogates (4,4′-dibromooctafluorobiphenyl,
deuterated PAHs, PCB30, PCB103, PCB207) and triple extracted by ASEwith dichloromethane. A percentage lipid analysis was performed on analiquot of the fish sample extract by oven drying (40 °C) overnight andundertaking gravimetric analysis. The remaining fish extracts andsediments were passed through a mini clean-up column incorporatingsodium sulphate, silica and alumina, to remove polar interferences. Lipidwas removed from fish samples using a Phenomenex Phenogel(300×21.2 mm, 10 μm and 300×21.2 mm, 5 μm in series) gel perme-ation column. Finally, all samples were fractionated on a silica columnusing pentane (F1) and 1:1 dichloromethane:pentane (F2). F1 containedall PCBs and the majority of OCPs. Sediment sample extracts were alsotreated with activated copper chips to remove sulphur. Internalstandards (tetrachlorometaxylene and octachlorostyrene) were addedto all extracts prior to analysis by GC. Quantitative analysis of PCBs andOCPs was carried out by capillary gas chromatography using a massselective detector in selected ion mode (GC–MS–SIM), on an Agilent6890 GC with 5975B MSD in splitless injection mode, using a30 m×0.25 mm i.d. DB-5 ms GC column with helium carrier gas. SIMions used are contained in supplementary information. Final concentra-tions have been corrected for surrogate recoveries, with detection limitsfor OCPs and PCBs ranging from 0.05 to 0.2 ng/g and 0.1 to 0.3 ng/g dryweight, respectively. Checks on method performance incorporatedthe analysis of in-house reference standards, standard referencematerials, blanks and GC check standards. This information is availableon request.
The analysis of metals in fish, watercress and sediment sampleswas carried out by a commercial laboratory (Hill Laboratories, 2010),following established procedures involving acid digestion and analysisby ICP-MS.
Fish and sediment samples were analysed for a range ofOCPs including DDT and DDT metabolites, chlordanes andchlordane metabolites, hexachlorobenzene (HCB), lindane (γ-hexachlorocyclohexane; γ-HCH) and dieldrin. ΣDDT is definedas the contribution from p,p′-DDT, o,p′-DDT, p,p′-DDE, o,p′-DDE,p,p′-DDD and o,p′-DDD. ΣChlordanes is defined as the contri-bution from cis and trans nonachlor, cis and trans chlordane,heptachlor, and cis and trans heptachlor epoxide. All sampleswere analysed for eight selected heavy metals; arsenic (As),cadmium (Cd), chromium (Cr), copper (Cu), lead (Pb), mercury
Table 1Contaminant data and risk values used in risk assessments for biological speciescollected from South Canterbury, New Zealand (2010).a
Species Compound Contaminantconcentration(μg/kg wet weight)
Watercress Cadmium 8.1 17.7 NA 80 1.0E−03Mercury 0.0 0.7 NA 80 1.0E−04Arsenic 12.4 31.8 1.5 80 3.0E−04Zinc 2657 4815 NA 80 3.0E−01Nickel 76.6 189.8 NA 80 2.0E−02Chromium 43.0 54.0 NA 80 3.0E−03
a Local consumption rates are species specific with median consumption of 6.1, 4.0,4.7 and 6.0 g/day for eels, trout, flounder and watercress respectively.
b CSF = cancer slope factor; BW = body weight, RfD = reference dose, NA = notapplicable, and ND = not determined.
c Arsenic risk calculation subsequently reduced by a factor of 10 for risk assessmentto reflect an approximate inorganic portion of total arsenic of 10% and provide aprotective estimate of health risk (US EPA, 2003).
2032 M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
(Hg), nickel (Ni) and zinc (Zn). Eel tissues were also analysedfor PCBs (32 congeners ranging from PCB 8 to PCB 209 withtotal PCBs or 32ΣPCB defined as sum of 32 congeners, seeSupplementary information for individual congeners). Water-cress was analysed for eight heavy metals only as describedabove. All analyte concentrations were initially reported on adry weight basis (see Supplementary information Tables S6–S11). Moisture content, calculated from original dry weight/wetweight data, was used for subsequent wet weight corrections ofanalytical data (Supplementary information; Tables S2–S4). Thisis required for comparisons of contaminant data with literatureand to enable calculation of appropriate risk factors based onwet weight consumption levels.
2.5. Risk assessment
Risk assessment calculations were undertaken for both cancer andnon-cancer risks. Cancer oral slope factor (CSF) and reference doses(RfD) for chronic non-cancer oral exposure were obtained from USEPA Integrated Risk Information System (IRIS) (US EPA, 2010), withthe exception of CSF and RfD for PCBs which were based on US EPAguidelines (US EPA, 2000). For mercury the RfD for methylmercurywas used (US EPA, 2010). No CSF or RfD values were available fromeither source for lead or copper. The US EPA concluded that it wasinappropriate to develop an RfD for lead. Although the US EPA hasclassed lead as a probable human carcinogen, it does not provide a CSFfor quantitative evaluation, due to many uncertainties in quantifyingthe risk (US EPA, 2010). Lindane was not detected in any sample sowas removed from the risk assessment calculations. We calculatedthreshold “trigger” values based on rearranged US EPA formulae (USEPA, 2000) which has been outlined previously (Watanabe et al.,2003). Contaminant-specific lifetime excess cancer risks werecalculated and from these data, the cumulative carcinogenic riskwas determined by summing all individual cancer risks. A margin ofexposure (MOE) was calculated to evaluate non-carcinogenic chronicrisk from individual contaminants associated with each species. Anarbitrary “risk flag” of 1 was set for MOE and 10−5 for lifetime cancerrisk (see discussion).
Data were analysed individually for each of the three fish speciescollected and an associated cumulative risk assessment carried out foreach contaminant (organochlorines and heavy metals) within thatspecies. Watercress was treated separately, with a risk assessment forheavy metals only being performed, since significant bioaccumulationof organic contaminants would not occur in aquatic vegetation.
For the purpose of this risk assessment, all analyte concentrationsbelow detectable concentrations were substituted with a value ofzero. The median and 95th percentile values for each contaminantacross all sites sampled was calculated (Table 1) to provide a measureof the range of potential risks, with the 95th percentile representing a“worst case” estimate for this assessment.
We calculated indicative local consumption rates based onArowhenua survey participants. However, as we had a limited surveysize (n=12), it was also necessary to include other consumptionscenarios for comparison. For fish, four different scenarios were used;species-specific locally-derived rates, a New Zealand average con-sumption (32 g/day), high consumption (43 g/day) and high energydiet (66 g/day) (Kim and Smith, 2006). Information onwatercress as afood has been summarised by Turner et al. (2005). Russell et al.(1999) identified that consumption of watercress at least once perweek was reported by 14% of Māori respondents, 13% were PacificIsland respondents and 1% were New Zealand European or otherethnic groups, with an average serving size for those consuming of230 g. Different watercress consumption rates were used, comprisinglocally-derived values plus low (7.1 g/day; equivalent to one servingof 230 g/month), medium (33 g/day; equivalent to one serving of230 g/week) and high (164 g/day; equivalent to five servings of
230 g/week) consumption rates. For consistency, an average bodyweight of 80 kg (Kim and Smith, 2006) was also used.
3. Results and discussion
Median and 95th percentile wet weight values were calculated fororganochlorines and the heavy metals cadmium, mercury, arsenic,zinc, nickel and chromium, for each species of fish and for watercress(Table 1). The median value was chosen over an arithmetic mean toremove the large influence of contaminant outliers in a relatively smallsample size and is used to determine what likely contamination loadwould be expected if harvestingwere to occur randomly across all sites
Fig. 2. Concentrations (ng/g wet weight) of DDT, DDT metabolites and isomers in biotaor sediment from sampling sites in South Canterbury, New Zealand. A (eels); B (troutand flounder); and C (sediment). See Fig. 1 for location of sampling sites. Note differingaxis scales.
2033M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
studied. The 95th percentile is a representative of a “worse case”scenario if harvesting was to occur of only the most contaminated kai.
Biometric data for each species is shown in Supplementaryinformation Tables S2–S4. Of particular note, eel measurements were;length 385–700 mm (median 598 mm), weight 123–862 g (median468 g), age 11–32 years (median 16 years), lipid 3.7–39.5% (median18.4%) (Supplementary information Table S2). The data for the onelong-fin eel collected (Ohapi Creek)was incorporatedwith the short-fineel data.
3.1. Organochlorine pesticides
Total DDT concentrations (ΣDDT) and the proportions of DDTisomers and metabolites are presented for eels (Fig. 2A), trout andflounder combined (Fig. 2B) and sediment (Fig. 2C). Also shown ineach figure are the median and 95th percentile values across all thesites.
The ΣDDT concentrations for eels ranged from 9.0 to 377 ng/g(ww) with a median of 32.0 ng/g and 95th percentile of 313 ng/g(Fig. 2A). Three sites had markedly higher ΣDDT concentrations foreels, namely Winchester (214 ng/g), Ohapi Creek (236 ng/g) andDoncaster (377 ng/g).Winchester andOhapi Creek are predominantlyrural areas, while Doncaster is part of the Washdyke industrial area,adjacent to the major town in the region, Timaru. Due to a limitedsample size, and for the purposes of this discussion, trout and flounderhave been grouped together. ΣDDT concentrations for trout andflounderwere in the range 2.3 to 35.7 ng/g, with amedian of 14.2 ng/gand 95th percentile of 29.2 ng/g (Fig. 2B). The concentrations of ΣDDTin trout and flounderwere generallymuch lower than for eelswith thehighest concentration (35.7 ng/g; Washdyke Lagoon) approximately10-fold less than the highest concentration for eels (377 ng/g;Doncaster). However, as shown in Fig. 2A and B, with the exceptionof the three highly contaminated eel samples alluded to earlier, mosteel ΣDDT concentrations were well below 50 ng/g and morerepresentative of those concentrations found in trout and flounder.Indeed,when the three sites ofWinchester, Ohapi Creek andDoncasterare removed from calculations, the adjusted median ΣDDT concen-tration for eels is 19.4 ng/g, not dissimilar from the trout and floundermedian value of 14.2 ng/g.
Washdyke Lagoon had the highest ΣDDT concentration forflounder (35.7 ng/g), whereas a relatively low eel concentration(19.4 ng/g) was measured at this site. The eel collected fromWashdyke Lagoon was relatively small; its length of 385 mm, weightof 123 g and lipid content of 3.7%, was very low compared with theregional median values of 598 mm, 468 g and 18.4% respectively (seeSupplementary information; Table S2). The low lipid content recordedfor this eel provides some explanation for the lower than expectedΣDDT concentration recorded due to the reduced bioaccumulation oforganic contaminants.
The highest concentrations of ΣDDT in the biota were generallyconcordant with the sediment concentrations (Fig. 2C), with OhapiCreek, Doncaster and Washdyke Lagoon having concentrations of 7.0,26.5 and25.7 ng/g respectively, compared to themedianof 1.0 ng/g. Theexceptions were Winchester, which showed high eel concentrationsand elevated trout tissue concentrations but low sediment concentra-tions, and the low bioaccumulation in the eel at Washdyke, as notedabove.
The percentage of DDT to ΣDDT in fish was on average 9% andwithin a range of 1.0% to 38.2% (see Supplementary information; TableS6), suggesting, in most cases, limited or no fresh input of DDT to theenvironment. The main exception to this was Washdyke Lagoon, witha percentage of 38.2% and 13.8% for eel and flounder respectively.Washdyke Lagoon is in an industrial area but is not a catchment for themajor rivers in the area.
Other organochlorine pesticides were either undetected or detectedin much lower concentrations than any of the DDT congeners
(Supplementary information; Table S6). Dieldrin wet weight correctedconcentrations ranged from b0.05 to 16.3 ng/g in eels and 0.08 to0.65 ng/g in trout and flounder. The most contaminated sites, with eelconcentrations over 1 ng/g were Doncaster (16.3 ng/g), Ohapi Creek(3.8 ng/g) and Winchester (2.2 ng/g).
Table 2Comparisons of organochlorine concentrations in eels and trout between this study(2010) and the New Zealand National Survey (Buckland et al., 1998a).
Species Compound Arowhenua (2009) MfE organochlorinesprogramme (1998)a
2034 M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
Total chlordane (ΣChlordanes) ranged from b0.1 to 10.6 ng/g (ww)in eels and b0.1 to 0.4 ng/g (ww) in trout and flounder. Doncaster(10.6 ng/g, eel) and Winchester (1.5 ng/g, eel) were the two mostcontaminated sites, with all other concentrations well below 1 ng/g.
Hexachlorobenzene wet weight concentrations ranged from b0.02to 0.44 ng/g in eels and b0.02 to 0.03 ng/g in trout and flounder, withthe most contaminated sites being Doncaster (0.44 ng/g, eel),Winchester (0.24 ng/g, eel) and Waihi River (0.22 ng/g, eel). Lindane(γ-hexachlorocyclohexane, (γ-HCH)) was not detected in any biotasample (limit of detection 0.04–0.08 ng/g, ww).
3.2. PCBs
PCBswere analysed in eels only,with a total of 32 congeners (32ΣPCB)included in the PCB suite (see Supplementary information; Table S8).Total wet weight concentrations ranged from 0.5 to 58.3 ng/g (Fig. 3),with the most elevated concentrations found at Doncaster andWinchester (58.3 and 22.7 ng/g respectively). PCBs were nevermanufactured in New Zealand, but were imported and used extensivelyin the electricity industry as insulating fluids or resins in transformersand capacitors (Bucklandet al., 1998b). AsDoncaster is an industrial site,high relative concentrations of PCBs are not a surprising result, howeverWinchester is a small rural town, so elevated concentrations at thislocation (relative to other rural sites in the area) are an unexpectedfinding.
Individual PCB congeners can exhibit both ‘dioxin-like’ toxicity and‘non-dioxin-like’ chronic effects. Dioxin-like PCBs elicit toxicity bybinding to the aryl hydrocarbon receptor (AhR), of which there are 12congeners (Van den Berg et al., 2006). Six of these dioxin-likecongeners were part of the PCB suite analysed here, specifically 77,118, 105, 126, 156 and 169. The two congeners that have the highestWorld Health Organization Toxic Equivalency Factors (WHO-TEF)(World Health Organization, 2009) values and hence contribute mostto dioxin-like toxicity (126 and 169), were not detected in any of theeels, with limits of detection at 0.1 ng/g (dw). The Total ToxicEquivalency (TEQ) calculated using the WHO-TEF (2005) values foreels from our study are calculated using upper bound concentrations,where values for different congeners less than the limits ofquantitation are equal to the limit of quantitation. For congeners126 and 169, this value is the significant determinant for TEQ ascalculations with values of 0.1 ng/g afford a TEQ range of 4.88–5.19 pg/g (ww). All these values are within a narrow range and arenear the action level of the European Commission Recommendation2006/88/EC (European Commission, 2006) for dioxin-like PCBs of6.0 pg/g (ww). This suggests a limitation in the use of thismethodology where detection limits of congeners 126 and 169 aresufficiently high to not allow accurate calculation of TEQs andtherefore significantly bias the results.
0
10
20
30
40
50
60
Waih
i Rive
r
Winc
heste
r
Temuk
a
Te Nga
Wai
Opihi b
elow P
P
Ohapi
Creek
Orari
Ohapi
Opihi u
pstre
am
Donca
ster
Was
hdyk
e La
goon
Co
nce
ntr
atio
n (
ng
/g, w
et w
eig
ht)
Fig. 3. Concentrations (ng/g wet weight) of 32ΣPCB in eels from sampling sites in SouthCanterbury, New Zealand.
To allow comparisons of PCB contaminant concentrations in eels toother studies, the sum of 7 indicator PCBs (7ΣPCB; 28, 52, 101, 118, 138,153 and 180) are often chosen, which are those that are known tobiomagnify in the food web. 7ΣPCB wet weight eel concentrationsranged from 0.4 to 37.2 ng/g, with a median of 1.3 ng/g. Theseconcentrations are considerably lower than the sum of indicator PCBsfound recently in Poland (4.0–533.9 ng/g ww) (Szlinder-Richert et al.,2010) or Scotland (median values from 5.9 to 1878 ng/g ww)(Macgregor et al., 2010), but comparable with eels from Ireland(1.94–18.1 ng/g ww) (McHugh et al., 2010).
The New ZealandMinistry for the Environment (MfE) initiated theorganochlorines programme in 1995 and carried out a nationwideassessment of various organochlorine contaminants in urban soils,river water, eel and trout, concluding that New Zealand rivers are“relatively free of contamination” (Buckland et al., 1998a). One areathat was not part of this nationwide study was South Canterbury, theregion of this study and an area of extensive past and presentagriculture. Comparisons between organochlorine contamination ofeel and trout in the MfE study and this study are shown in Table 2. Foreels, the concentrations of the minor organochlorines were generallycomparable between the two studies, with the exception of trans-chlordane, which was detected in our samples at concentrations 5times higher than reported in the MfE study (1.21 vs 0.24 ng/g). Alsofound in higher concentrations were maximum levels of p,p′-DDD(60.2 vs 33.1 ng/g), p,p′-DDE (287 vs 155 ng/g) and total PCBs (58.3 vs18.5 ng/g). Interestingly, the opposite was true for trout, where themaximum contaminant concentrations in our study were equal to orlower than those found in the MfE study (Table 2).
Although the South Canterbury region contains some of thehighest eel contamination of organochlorines nationally (maximumconcentration of p,p′-DDE and total PCBs of 287 and 58.3 ng/grespectively), it is difficult to compare concentrations with overseasresults due to substantial variations in contamination found. PCBswere covered earlier in this paper, however p,p′-DDE (the dominantDDT breakdown product) was found in eels with maximum wetweight concentrations of 7.10 ng/g in Ireland (McHugh et al., 2010),28.0 ng/g (Storelli et al., 2007) and 88.0 ng/g (Ferrante et al., 2010) inItaly and 3422.6 ng/g in Belgium (Maes et al., 2008), to name a few
2035M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
recent studies. As such, it is worthwhile assessing the individual riskof each contaminant, which is addressed in Section 3.4.
3.3. Heavy metals
The analysis of eight heavy metals, arsenic (As), cadmium (Cd),chromium (Cr), copper (Cu), lead (Pb), mercury (Hg), nickel (Ni) andzinc (Zn), was carried out on all fish samples, an important localvegetable, watercress and sediment. Full data are presented in theSupplementary information (Tables S4–S6). Median, minimum andmaximum values are presented in Table 3, along with heavy metalcontamination profiles of fresh fish and silverbeet (a comparablevegetable) from the 2009 New Zealand total diet survey (NZTDS),which tested contamination concentrations of food from a range ofretail outlets in four major regions (NZFSA, 2009).
From this current study, mercury contamination was highest ineels with a median value of 0.34 μg/g (range 0.02–0.56 μg/g). The Hgconcentrations were much lower in trout (median concentration0.11 μg/g; range 0.05–0.50 μg/g), lower again in flounder (median0.04 μg/g; range 0.02–0.10 μg/g), and generally below detection inwatercress (Table 3). The NZTDS data had median values of 0.12 μg/gfor fresh fish, very close to that observed in trout from SouthCanterbury and generally below detection in silverbeet, consistentwith that seen in watercress. The mercury contamination in eels wasclose to – and at two sites exceeded – the Food Standards AustraliaNew Zealand (FSANZ) limits of 0.5 μg/g for most fish (including eeland trout) (FSANZ, 2009, 2010), but not the EC limit of 1 μg/g foreels (European Commission, 2006). The source of mercury in thisarea is unclear. Unlike parts of the North Island of New Zealand,South Canterbury does not have geothermal inputs to lakes andrivers, which contribute mercury, arsenic and other metals tofreshwaters.
The heavy metal contamination profile was different for arsenic,being undetectable in eels, highest concentrations in trout (median0.27 μg/g; range bLOD — 0.89 μg/g), intermediate concentrations inflounder (median 0.12 μg/g; range 0.06–0.56 μg/g) and low concentra-tions in watercress (median 0.01 μg/g; range bLOD — 0.03 μg/g). Thefish total arsenic (Astot) concentrations from this study were lowcompared to theNZTDS concentrations for fresh fish (median 3.69 μg/g;range 2.08–6.31 μg/g). The maximum Astot concentration (0.89 μg/g;trout Opihi river mouth) was well below the FSANZ standard of 2 μg/gfor fish and the maximum watercress Astot concentration (0.03 μg/g)was well below the FSANZ standard set for seaweed of 1 μg/g (FSANZ,2009). Furthermore, the FSANZ standards (for fish and seaweed) are setfor inorganic arsenic (Asi) only and, in the case of fish at least, theinorganic proportion of the total arsenic found is likely to be very low.The US EPA concluded that an assumption that 10% of Astot is Asi inresident freshwater fish provided a protective estimate of health risk(US EPA, 2003), which was supported by a more recent survey of theliterature, which concluded that for freshwater fish, Asi was 10% of Astotat the 75th percentile (Schoof and Yager, 2007). These data suggest thatthe As concentrations found in this study are well below the FSANZinorganic standard of 2 μg/g for fish.With no geothermal activity in thisregion, the arsenic concentrations in biota could be partly due to manyunknown contaminated sheep dip sites in the area.
Lead in fish had a maximum concentration of 0.08 μg/g (ww) ineels (Table 3), which was well below the FSANZ standard of 0.5 μg/g.Watercress lead concentrations ranged from 0.03 to 0.10 μg/g. The EChas set a standard of 0.3 μg/g for leaf vegetables (EuropeanCommission, 2006) which is consistent with the FSANZ standard forbrassicas (FSANZ, 2009), with the most contaminated watercresssample falling well within these standards.
Cadmium was virtually undetectable in all samples, with amaximum fish concentration of 0.02 μg/g (ww) in eels and maximumwatercress concentration of 0.02 μg/g (ww) (Table 3). The FSANZ doesnot have a maximum standard for cadmium in fish (FSANZ, 2009),
2036 M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
however the EC standard is 0.1 μg/g for eels and 0.05 μg/g for otherfish (European Commission, 2006). The concentrations of cadmium inthis study were well below these values. The FSANZ standard forcadmium in leafy vegetables is 0.1 μg/g (FSANZ, 2009), with thehighest concentration found in this study 5-fold lower than this.
No maximum standards have been set by FSANZ (2009) or the EC(2006) for copper, zinc, nickel or chromium, however zinc, nickel andchromium have been incorporated into the risk assessment for non-cancer risk only.
To benchmark the overall contamination, sediment heavy metalconcentrations for the sites where kai was harvested in this studywere compared with the Australian and New Zealand EnvironmentConservation Council (ANZECC) Interim Sediment Quality Guidelines(ISQG)(ANZECC, 2000) (Table 3). Low and high ISQG have been set byANZECC, corresponding to the effects range-low and effects range-median adapted from Long et al. (1995). These sediment guidelineswere only exceeded on one occasion. The low ISQG value of 200 μg/gfor zinc was exceeded at Doncaster, with a value of 220 μg/g (seeSupplementary information, Table S11).
The use of guideline values is only one way of determining risk.There is added benefit from considering specific consumption rates ofa food (in this case wild caught kai). Furthermore, additive risk is notcovered by contaminant specific maximum residue limits and for thisa more detailed risk assessment is needed.
3.4. Risk assessment
Indicative local average consumption rates of wild kai by Māoriwere calculated as 6.1, 4.0 and 4.7 g/day for eels, trout and flounderrespectively (Tipa et al., 2010a). Watercress consumption wascalculated at 6.0 g/day (Stewart et al., 2010). These consumption
Table 4Risk assessment calculations of lifetime cancer risk for four separate consumption rate scen
Cancer risk values above threshold of 10−5 are bold text. Median and 95th percentile contamscenarios are described in Materials and methods section.
a Arsenic risk calculation reduced by a factor of 10 to reflect an approximate inorganic porti
rates are markedly lower than average New Zealand consumptionrates for total fish (32 g/day) and watercress (33 g/day) and reflectthe generally supplementary dietary intake of wild kai. Even themaximum local consumption rates of 20.0, 13.3 and 13.3 g/day foreels, trout and flounder respectively are still well below the averageNew Zealand consumption rates for total fish. In contrast, the averagetotal fish consumption from the survey of local Māori was 43 g/day,putting local consumption rates into the New Zealand high categoryand highlighting that wild caught kai is only a small proportion of themain source of food for the local community.
Given that thederivationof local consumption rateswasbasedonsucha small sample size, scenarios of higher than average fish consumptionwere also included to ascertain the potential range of contaminantconcentrations that may present a risk to wild kai consumers. As such,NewZealandaverage, high andhigh energydietfish consumption rates of32, 43 and 66 g/day respectively and low, medium and high watercressconsumption rates of 7.1, 33 and 164 g/day respectively were used in therisk assessment process. Furthermore, as noted earlier, an adjustmentfactor of 0.1 was applied to total arsenic (Astot) concentrations in fish tobetter reflect the proportion of inorganic arsenic (Asi). As a protectivemeasure, watercress arsenic concentrationswere not adjusted for the riskassessment, as arsenic was assumed to be predominantly inorganic asobserved in some plants (Daus et al., 2005; Zhang et al., 2002).
For this screening assessment, contaminant specific lifetime excesscancer risks were calculated for four different consumption ratescenarios (Table 4) by following calculations used previously(Watanabe et al., 2003), specifically:
ination values in table are those across all sites for a specific contaminant. Consumption
on of total arsenic of 10% and provide a protective estimate of health risk (US EPA, 2003).
2037M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
A total cumulative excess cancer risk was calculated by summingeach contaminant-specific excess cancer risk per species (Table 4).
An “acceptable” lifetime cancer risk level of 10−6 (1 in 1,000,000)is considered by some countries or institutions as negligible (WorldHealth Organization, 2009) and a level of 10−5 (1 in 100,000) is set byUS EPA in their “Guidance for Assessing Chemical Contaminant Datafor Use in Fish Advisories” (US EPA, 2000). As such, we set a 10−5
lifetime cancer risk level as a threshold value to determine potentialrisks of individual and/or total contaminants from our fish andwatercress samples. Any lifetime cancer risk level exceeding 10−5 isindicated in bold text (Table 4).
The data in Table 4 indicate that for a median contaminationprofile, i.e., where harvesting is carried out at all sites randomly, thenthe only individual contaminant excess cancer risk exists for arsenic.In trout, arsenic is shown to be a risk at a New Zealand averageconsumption rate (32 g/day) and for flounder this risk is onlyapparent for New Zealand high energy diets (66 g/day) and above.However, if all contaminant excess cancer risks are summed to give atotal excess cancer risk, then a potential risk is apparent for all thesampled fish consumed at New Zealand high consumption rates(43 g/day) and trout and flounder consumed with New Zealandaverage consumption rates (32 g/day).
Table 5Risk assessment margin of exposure calculations of chronic non-cancer risk for four separa
Margin of exposure (MOE) values above threshold of 1 are bolded. Median and 95th percenConsumption scenarios are described in Materials and methods section.
a Arsenic risk calculation reduced by a factor of 10 to reflect an approximate inorganic porti
When using the 95th percentile data to approximate a scenario ofharvesting of themost contaminated kai only, total excess cancer risksareflagged across all consumption rates for eels and fromaverageNewZealand consumption rates and above for trout and flounder. For eels,the data (Table 4) highlights dieldrin as having the highest individualrisk, followed closely by total PCBs and p,p′-DDE, which shows thatthere is significant risk if the most contaminated eels were predom-inantly eaten and at local consumption rates. Doncaster, Ohapi Creekand Winchester were clearly identified as the sites where elevatedconcentrations of contaminants from all three organochlorine groupsstated above were recorded. Doncaster was the most contaminated,followed by Ohapi Creek then Winchester.
Excess lifetime cancer risk from watercress was evaluated onarsenic only and was generally of minimal concern (Table 4).However, arsenic is still an apparent risk for high consumption rates(164 g/day) when the median contamination is considered, or whenmedium and high consumption rates (33 and 164 g/day respectively)were used in combination with the most contaminated samples (95thpercentile). Watercress can be considered a low excess cancer riskbecause local median consumption rates (6.0 g/day) are much lower.
Using a cumulative risk assessment for excess cancer risk is usefulin that it gives a total risk of consuming a particular species, with the
te consumption rate scenarios for eels and fish from South Canterbury, New Zealand.
high energy This study NZ average NZ high NZ high energy
tile contamination values in table are those across all sites for a specific contaminant.
on of total arsenic of 10% and provide a protective estimate of health risk (US EPA, 2003).
2038 M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
proviso being that all known carcinogenic contaminants that arepresent in the food are accounted for and the analytical data are bothaccurate and sensitive. For example, PCBs were not analysed in troutor flounder, however with the risks revealed by PCBs in eels, thiswould be a prudent future analysis. Furthermore, for reasons of cost,dioxins were not analysed in any kai species. There could be potentialhealth implications, at least at some of the sites, e.g., there is a knownhistoric timber treatment facility around Temuka (Brett Mongilo,Environment Canterbury, pers. comm. March 2009). In addition,elevated contaminant risks for individual contaminants (i.e., arsenicin trout and flounder and certain organochlorines in eels) identifywhere a further, more refined, effort is needed in establishing theextent of this contamination in the region.
Contaminant specific chronic non-cancer risks were also calculatedfor all four consumption rate scenarios (Table 5), where a margin ofexposure (MOE) was calculated as for Watanabe et al. (2003):
An MOE greater than 1 indicates exposure to contamination greaterthan the safe dose for chronic non-carcinogenic effects and any valueover 1 was flagged (Table 5). For eels, the most significant non-cancerrisk was due tomercury contamination. Any consumption rate equal toor above the New Zealand average shows significant risk, however, therisk threshold was not exceeded for any species for the indicative localconsumption rates (Table 5). The difference between median and 95thpercentile contamination risk formercury is less than 2-fold, suggestingthat mercury contamination in eels from our study is reasonablyconsistent. This is in contrast to trout, where median contamination ofmercury does not yield aMOEover 1, but the95th percentile data showsrisk for the New Zealand average consumption rate or above. There isapproximately a 4-fold difference between the median and 95thpercentile mercury contaminant concentrations in trout. Flounder islargely a low non-cancer risk, even when collected from the mostcontaminated site(s) (95th percentile) with a high energy diet.
As stated previously, the FSANZ standard for mercury in most fish is0.5 mg/kg (FSANZ, 2009, 2010). FSANZuse aprovisional tolerableweeklyintake (PTWI) of 1.6 μg/kg body weight/week taken from the JointFAO/WHO Expert Group on Food Additives (JECFA), stating that intakesof up to about two times higher than the existing PTWI would not poseany risk of neurotoxicity in adults (World Health Organization, 2006).The PTWI of 1.6 μg/kg body weight/week equates to 2.3×10−4 mg/kg/day, which is a less conservative estimate of risk than the US EPA RfD of1×10−4 mg/kg/day (US EPA, 2010).
PCBs show a significant non-cancer risk for high and highenergy consumption rates of eels from the most contaminated(95th percentile) sites, with the New Zealand average consump-tion rate providing a MOE of 0.9686, just below the threshold(Table 5).
Although the high energy diet (66 g/day) appears to be anexcessive amount of fish consumption, this value is equivalent tothe estimated daily intake for adolescent males of 64.5 g/day (Smithand Lopipero, 2001). Consistent high consumption would not seemlike a realistic consumption rate of wild caught fish for an adolescentmale, but does illustrate what the consequences could be for any localindividual who has a fish diet that was of this level and predominantlysourced from the wild.
4. Conclusions
This study investigated the concentrations of bioaccumulativecontaminants in three species of fish (eel, trout and flounder) andone commonly harvested plant (watercress) from historic andcurrent kai collection sites throughout the rohe (tribal territory) ofArowhenua, in the region of South Canterbury, New Zealand.
Concentrations measured in this study were compared with thosefound elsewhere throughout the world and also with nationallyderived information.
Eels generally had higher concentrations of DDTs, dieldrin and PCBsthan other regions of New Zealand (Buckland et al., 1998a). Mercuryconcentrations exceeded the FSANZmaximum standards of 0.5 μg/g forfishat two sites butwere still below theECmaximumstandardof 1 μg/g,for eels. Trout and flounder had lower concentrations of organochlorinepesticides than eels.
Sediment heavy metal concentrations were generally belowANZECC-ISQ Low guidelines (ANZECC, 2000).
The lifetime cancer risk assessment identified organochlorines(dieldrin, PCBs and p,p′-DDE) in eels as primary contaminants ofconcern, while lifetime cancer risk from ingesting trout and flounderwas dominated by arsenic. The non-cancer chronic health risk wasdominated by mercury in eels and trout and PCBs were also shown tobe a risk in themost contaminated eels, based upon high consumptionlevels. The area of Washdyke consistently showed the highest risk ofconsumption of kai, however significant risk of consumption was notlimited to this area.
Indicative survey data showed that local consumption rates of wildcaught kai are lower than the average national consumption rates forfish or vegetables. The survey information also highlighted howgathering of wild kai has decreased from subsistence gathering duringhistoric times to a rarer sporadic activity, often undertaken for specialoccasions. However, even with low consumption rates of wild caughtkai, human health may still be at risk, especially if harvesting of kaiwas carried out consistently at one or more of the most highlycontaminated sites.
Further work is needed to accurately quantify the dietary exposurerisk of contaminants in freshwater and marine wild caught kai in thisregion. To achieve this, future work should include collecting multiplekai specimens from a larger range of sites (including nearshoremarine environments), expanding the contaminant dataset to includePCB analyses in all fish, pentachlorophenol (PCP) and dioxin (bothpotentially derived from historic timber treatment operations)analyses in areas of potential concern and metal speciation studieson arsenic and mercury. It is also necessary to obtain a more robustdataset of wild caught kai and commercial seafood consumption inthe region, by including larger numbers of local Māori and non-Māoriconsumers of wild kai in the questionnaire process and conducting arisk assessment for total fish diet which incorporates both wild andcommercial dietary consumption.
This study has determined the potential health risks for Māori whoconsume key species of locally caught freshwater and estuarine kaifrom in this region and identified the contaminants of concern. Thequantitative risk assessment provides the basis for establishingrecommended dietary consumption limits for harvest sites withinthe study region.
Acknowledgments
The authors thank Lindsay Hawke and Julian Sykes (NIWAChristchurch) for assistance with field sampling. We are grateful tothose members of Te Runanga o Arowhenua who contributed theirknowledge and participated in the interviews and questionnaires.This research was funded by the Health Research Council of NewZealand, contract HRC/207. Kai consumption data were collectedunder Ethics Approval # MEC/07/07/088.
Appendix A. Supplementary data
Supplementary data to this article can be found online atdoi:10.1016/j.scitotenv.2011.02.028.
2039M. Stewart et al. / Science of the Total Environment 409 (2011) 2029–2039
References
ANZECC. Australian and New Zealand guidelines for fresh and marine water quality.National Water Quality Management Strategy Paper No. 4. Canberra, Australia:Australian and New Zealand Environment and Conservation Council and Agricultureand Resource Management Council of Australia and New Zealand; 2000.
Buckland SJ, Jones PD, Ellis HK, Salter RT. Organochlorines in New Zealand: ambientconcentrations of selected organochlorines in rivers. Wellington: New ZealandMinistry for the Environment; 1998a.
Buckland SJ, Scobie S, Heslop V. Concentrations of PCDDs, PCDFs and PCBs in retail foodsand an assessment of dietary intake for New Zealanders. Wellington: New ZealandMinistry for the Environment; 1998b.
Butler CA, Timperley MH. Fertilised farmland as a source of cadmium in oysters. SciTotal Environ 1996;181:31–44.
Daus B, Wennrich R, Morgenstern P, Weiß H, Leonhardt Palmieri HE, Nalini HA, et al.Arsenic speciation in plant samples from the iron quadrangle, Minas Gerais, Brazil.Microchim Acta 2005;151:175–80.
Environment Canterbury. Opihi river regional plan. R00/16. Christchurch: ECan; 2000.Environment Canterbury. Managing old sheep dip and footbath sites. http://ecan.govt.
European Commission. Commission Regulation (EC) No 1881/2006 of 19 December2006 setting maximum levels for certain contaminants in foodstuffs; 2006.
Ferrante MC, Clausi MT, Meli R, Fusco G, Naccari C, Lucisano A. Polychlorinatedbiphenyls and organochlorine pesticides in European eel (Anguilla anguilla) fromthe Garigliano River (Campania region, Italy). Chemosphere 2010;78:709–16.
FSANZ. Standard 1.4.1 — contaminants and natural toxicants. V109; 2009.FSANZ. Standard 1.4.2 — Schedule 4. Foods and classes of foods. V119; 2010.Hill Laboratories. http://www.hill-laboratories.com/ (Accessed 3rd August, 2010).Hiroa TR. Maori food-supplies of Lake Rotorua, with methods of obtaining them, and
usages and customs appertaining thereto. Trans R Soc NZ 1921;53:433–51.Hoekstra PF, O'Hara TM, Backus SM, Hanns C, Muir DCG. Concentrations of persistent
organochlorine contaminants in bowhead whale tissues and other biota fromnorthern Alaska: implications for human exposure from a subsistence diet. EnvironRes 2005;98:329–40.
Johansen P, Muir D, Asmund G, Riget F. Human exposure to contaminants in thetraditional Greenland diet. Sci Total Environ 2004;331:189–206.
Kim N, Smith J. Review of science relating to discharges from the Kinleith pulp andpaper mill. Environment Waikato Technical Report. Environment WaikatoTechnical Report 2005/58. Hamilton: Environment Waikato; 2006.
Long E, Macdonald D, Smith S, Calder F. Incidence of adverse biological effects withinranges of chemical concentrations in marine and estuarine sediments. EnvironManage 1995;19:81–97.
Macgregor K, Oliver IW, Harris L, Ridgway IM. Persistent organic pollutants (PCB, DDT,HCH, HCB and BDE) in eels (Anguilla anguilla) in Scotland: current levels andtemporal trends. Environ Pollut 2010;158:2402–11.
Maes J, BelpaireC,GoemansG. Spatialvariationsand temporal trendsbetween1994and2005in polychlorinated biphenyls, organochlorine pesticides and heavy metals in Europeaneel (Anguilla anguilla L.) in Flanders, Belgium. Environ Pollut 2008;153:223–37.
McDowall RM. The reed field guide to New Zealand freshwater fishes. Auckland,New Zealand: Reed; 2000.
McHugh B, Poole R, Corcoran J, Anninou P, Boyle B, Joyce E, et al. The occurrence ofpersistent chlorinated and brominated organic contaminants in the European eel(Anguilla anguilla) in Irish waters. Chemosphere 2010;79:305–13.
MfE. Identifying, investigating and managing risks associated with former sheep-dipsites: a guide for local authorities. . ME 775Wellington: New Zealand Ministry forthe Environment; 2006.
NOAA. Sampling and analytical methods of the national status and trends programnational benthic surveillance and mussel watch projects 1984–1992. Technicalmemorandum NOS ORCA 71, III and IV; 1993.
NZFSA. Review of non-commercial wild food in New Zealand. NZFSA Draft PositionPaper (No: 01/05). Wellington: New Zealand Food Safety Authority; 2005.
NZFSA. 2009 New Zealand total diet study. Wellington: New Zealand Food SafetyAuthority; 2009.
Odland JO, Deutch B, Hansen JC, Burkow IC. The importance of diet on exposure to andeffects of persistent organic pollutants on human health in the Arctic. Acta Paediatr2003;92:1255–66.
Robinson BH, Duwig C, Bolan N, Kannathasan M, Saravanan A. Uptake of arsenic byNew Zealand watercress (Lepidium sativum). Sci Total Environ 2003;301:67–73.
Russell D, Parnell W, Wilson N. NZ Food: NZ People. Key results of the 1997 NationalNutrition Survey; 1999.
Schoof RA, Yager JW. Variation of total and speciated arsenic in commonly consumedfish and seafood. Hum Ecol Risk Assess 2007;13:946–65.
Smith A, Lopipero P. Evaluation of the toxicity of dioxins and dioxin-like PCBs: a healthrisk appraisal for the New Zealand population. Wellington: New Zealand Ministryfor the Environment; 2001.
Statistics New Zealand. 2006 Census. http://www.stats.govt.nz/Census/2006CensusHome-Page/ (Accessed 1st November, 2010).
Stewart M, Olsen G, Phillips N, Hickey C. Contaminants in Kai — Arowhenua rohe. Part 1:data report. HAM2010-105. Hamilton: National Institute of Water & AtmosphericResearch; 2010.
Storelli MM, Barone G, Garofalo R, Marcotrigiano GO. Metals and organochlorinecompounds in eel (Anguilla anguilla) from the Lesina lagoon, Adriatic Sea (Italy).Food Chem 2007;100:1337–41.
Szlinder-Richert J, Usydus Z, Pelczarski W. Organochlorine pollutants in European eel(Anguilla anguilla L.) from Poland. Chemosphere 2010;80:93–9.
Taylor R, Smith I, Cochrane P, Stephenson B, Gibbs N. The State of New Zealand'sEnvironment 1997. Wellington: New Zealand Ministry for the Environment; 1997.
Tipa G, Nelson K, Downs S, Home M, Phillips N. A survey of wild kai consumption in theArowhenua rohe. HAM2010-098Hamilton:National Institute ofWater&AtmosphericResearch; 2010a.
Tipa G, Nelson K, EmeryW, Smith H, Phillips N. A survey of wild kai consumption in theTe Arawa rohe. HAM2010-096Hamilton: National Institute ofWater & AtmosphericResearch; 2010b.
Turner N, Cressey P, Lake R, Whyte R. Review of non-commercial wild food inNew Zealand. Wellington: ESR report for New Zealand food safety authority; 2005.
US EPA. Method 600/4-81-055. Interim methods for the sampling and analysis orpriority pollutants in sediments and fish tissue. Revised October 1980 (NERL);1977.
US EPA. Method 503/6-90-004. Analytical methods for US EPA priority pollutants and301(h) pesticides in estuarine andmarine sediments. US Environmental ProtectionAgency; 1986.
US EPA. Guidance for assessing chemical contaminant data for use in fish advisories.Volume 2: Risk assessment and fish consumption limits — third edition. USEnvironmental Protection Agency; 2000.
US EPA. Columbia river basin fish contaminant survey. US Environmental ProtectionAgency; 2003.
US EPA. Integrated risk information system. http://www.epa.gov/ncea/iris/index.html(Accessed 5th July, 2010).
Van den Berg M, Birnbaum LS, Denison M, De Vito M, Farland W, Feeley M, et al. The2005 World Health Organization reevaluation of human and mammalian toxicequivalency factors for dioxins and dioxin-like compounds. Toxicol Sci 2006;93:223–41.
Van Oostdam J, Gilman UA, Dewailly E, Usher P, Wheatley B, Kuhnlein H, et al. Humanhealth implications of environmental contaminants in Arctic Canada: a review. SciTotal Environ 1999;230:1-82.
Van Oostdam J, Donaldson S, Feeley M, Tremblay N. Canadian arctic contaminantsassessment report II: human health. Ottawa: Northern Contaminants Program;2003. p. 127.
Watanabe KH, Desimone FW, Thiyagarajah A, Hartley WR, Hindrichs AE. Fish tissuequality in the lower Mississippi River and health risks from fish consumption. SciTotal Environ 2003;302:109–26.
World Health Organization. WHO Technical Report Series 940. Evaluation of certainfood additives and contaminants; 2006. Geneva.
World Health Organization. Environmental health criteria 239. Principles for modellingdose-response for the risk assessment of chemicals. World Health Organisation(WHO) International Programme on Chemical Safety (IPCS), Geneva. Environmentalhealth criteria 239. Geneva: World Health Organisation (WHO) InternationalProgramme on Chemical Safety (IPCS); 2009.
Zhang W, Cai Y, Tu C, Ma LQ. Arsenic speciation and distribution in an arsenichyperaccumulating plant. Sci Total Environ 2002;300:167–77.
