Jens Søndergaard

Gert Asmund

Poul Johansen

Frank Rigét

Long-term response of an arctic fiord system to lead-zinc mining and submarine disposal of mine waste (Maarmorilik, West Greenland)

Keywords: Mining, Submarine tailings disposal, Pollution monitoring, Heavy metals, Lead, Zinc, Mytilus edulis, Fucus vesiculosus, Greenland, Arctic

Contamination by lead (Pb) and zinc (Zn) was studied in seawater, sediment, seaweed and blue mussels near the former "Black Angel Pb-Zn Mine" in Maarmorilik, West Greenland. The mine was operated during the period 1973-90 when mine waste (tailings and later waste rock) was discharged directly into the sea. Metal concentrations peaked during the mining period and Pb and Zn in seawater within the discharge area were measured up to 440 and 791 µg L -1 , respectively. Pb concentrations in sediment, seaweed and mussels just outside the discharge area were measured up to 187, 84 and 2,650 and Zn up to 655, 363 and 1,190 µg g -1 dry wt., respectively. Within the discharge area, seawater metal concentrations (especially Pb) decreased abruptly after mine closure.

Metals concentrations in sediment and biota, however, decrease more slowly and two decades after mine closure seaweed and blue mussels are still contaminated 12 km from the mine.

Introduction

The Black Angel Lead-Zinc Mine in Maarmorilik in West Greenland (Figure 1) is an important example in mining history where discharge of waste materials into the sea caused a significant release and dispersion of metals within a larger fiord system (Asmund, 1992a;[START_REF] Larsen | Lead and zinc in sediments and biota from Maarmorilik, West Greenland: an assessment of the environmental impact of mining wastes on an Arctic fjord system[END_REF][START_REF] Elberling | Geochemical trends in metalcontaminated fiord sediments near a former lead-zinc mine in West Greenland[END_REF]. The Black Angel Mine was operated between 1973 and 1990 and during this period, so-called mine tailings from ore treatment were discharged into a small partly-enclosed fiord Affarlikassaa (named the A-fiord). In the years after the tailings disposal began, elevated concentrations of metals including lead (Pb) and zinc (Zn) were found in seawater, sediment and biota, not only in the A-fiord but also in a larger area covering the outer fiord Qaamarujuk (Qfiord). Later, detailed geochemical studies revealed that the tailings contained significant amounts of Pb and Zn minerals that were soluble in the ambient seawater [START_REF] Poling | Importance of geochemistry: The Black Angel Lead-Zinc Mine, Greenland[END_REF]. Furthermore, seasonal destratification of the water body in the A-fiord was found to take place during winter/spring, which allowed for a complete mixing of surface water with metal-rich bottom water and subsequently a dispersal of contaminants over the sill between the A-fiord and the Qfiord [START_REF] Poling | Importance of geochemistry: The Black Angel Lead-Zinc Mine, Greenland[END_REF]. Other sources of contamination in Maarmorilik include dust from the ore-milling process and dumps of waste rock established on land below the mine entrances containing rocks with elevated levels of Pb and Zn. The largest and most contaminating of the land- Today (2010), there are ongoing plans to re-open the mine within the next few years and this study seeks to evaluate the present state of the environment in the A-and Q-fiords 20 years after mineclosure. This is important not only as an environmental "baseline study" for the re-opening of the mine but also as a study of long-term recovery of an arctic fiord system subjected to mining pollution, which may also provide a valuable reference for future cases of submarine tailings disposal. The study is based on analyses of seawater, sediment and the marine organisms -seaweed (Fucus vesiculosus/Fucus distichus) and blue mussels (Mytilus edulis). These relatively sedentary organisms concentrate metals, reflect different bioavailable metal-sources within the seawater and have been used as biological monitors of heavy metal pollution in a number of previous studies, including the so-called Mussel Watch Programs [START_REF] Rainbow | Biomonitoring of heavy metal availability in the marine environment[END_REF][START_REF] Cairrao | Fucus spp. as a mercury contamination bioindicator in coastal areas (Northwestern Portugal)[END_REF][START_REF] Connor | Trends in chemical concentrations in mussels and oysters collected along the US coast: Update to 2003[END_REF][START_REF] Torres | Biochemical biomarkers in algae and marine pollution: A review[END_REF]. Data include those from samples taken during the mining period before 1990 and until 2009 and make a detailed consistent record of high-resolution data from environmental monitoring and research in the area.

Site description

Maarmorilik is situated adjacent to the two fiords, Affarlikassaa and Qaamarujuk (A-and Q-fiords), in the inner part of the Uummannaq fiord complex in West Greenland (71°07' N; 51°15' W) (Figure 1). The closest community is Ukkussissat, 25 km to the west and Uummannaq, the main settlement in the area, is situated 80 km south of Maarmorilik. The climate is arctic with maximum summer temperatures around 10°C and minimum winter temperatures below -30°C. Winds are dominated by strong easterly winds coming from the Greenland Ice Sheet in the bottom of the Qfiord, which result in sparse precipitation in the area (<100 mm a year).

The A-fiord covers an area of 2 km 2 and has an average water depth of 30 m. The fiord is partly separated from the outer Q-fiord by a sill at 23 m depth. The A-and Q-fiords can be described as stratified estuaries during summer. The stratification is caused by melt water from the rivers that flow on top of the heavier seawater creating a distinct and roughly stagnant water mass in the bottom of the Q-fiord and A-fiord. The annual fresh water input to the Q-fiord equals approximately 90x10 6 m 3 from the river in the bottom of the A-fiord and 22x10 6 m 3 from the Wegener River in the inner part of the Q-fiord [START_REF] Møller | Hydrodynamics of an arctic fiord[END_REF]. In almost every winter, however, a complete vertical mixing of the stratified water bodies take place due to freezing of the surface water and water inflow over the sill between the A-fiord and Q-fiord [START_REF] Møller | Hydrodynamics of an arctic fiord[END_REF]. The annual seawater

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4 temperatures in the fiords range between -2 and 5°C, the salinity between 31 and 34 psu and the pH between 7.8 and 8.1. The tidal range in the fiords is around 1 m and sea ice, 0.5-1 m thick, usually develops in October and covers the fiords until May.

The geological Maarmorilik Formation includes many carbonate-hosted Pb-Zn ores, which are primarily located in the 'Black Angel Mountain', and the area is part of the Archean Foxe-Rinkian mobile belt complex of North-East Canada and Central West Greenland [START_REF] Escher | Rinkian mobile belt of West Greenland[END_REF]. The name "Black Angel" refers to a pelite outcrop forming a dark angel-like figure high on a 1100 m marble cliff face above the A-fiord. The massive ores, up to 30 m thick, consist of galena (PbS), sphalerite (ZnS) and pyrite (FeS 2 ) with accessory ore minerals such as pyrrhotite, chalcopyrite, tennantite and arsenopyrite. Prior to mining, the Black Angel deposit consisted of ten major ore bodies with a total of 13.6×10 6 tons containing 4.0% Pb, 12.3% Zn and 29 ppm Ag. Out of these, 11.2×10 6 tons were mined in the period 1973-90 by the company Greenex A/S [START_REF] Thomassen | The Black Angel lead-zinc mine at Maarmorilik in West Greenland[END_REF].

During the mining process, ore material was transported from the mine entrances at 600 m altitude by means of cable cars across the A-fiord to a floatation plant in Maarmorilik. Here, concentrates were produced, loaded onto ships and transported to smelters in Europe. For ore treatment, conventional rod/ball milling was applied to liberate Pb and Zn. This was followed by froth floatation to produce separate Pb and Zn concentrates plus a waste product, tailings [START_REF] Poling | Importance of geochemistry: The Black Angel Lead-Zinc Mine, Greenland[END_REF]. Permission for submarine disposal of tailings was given to the mining company and the tailings were subsequently discharged as a suspension in seawater and led directly into the Afiord at around 30 meters depth. In the 1970s, the annual tailings inputs from the flotation mill to the A-fiord were 4.5×10 5 t, which increased to 6×10 5 t in the 1980s. During the mining period it is estimated that more than 8×10 6 t of mine tailings containing more than 2.2×10 4 t Pb and 5×10 4 t Zn were discharged into the A-fiord [START_REF] Elberling | Geochemical trends in metalcontaminated fiord sediments near a former lead-zinc mine in West Greenland[END_REF]. As part of the mine exploitation, several waste rock dumps were established below the mine entrances in the Black Angel Mountain. One of the most polluting of these waste rock dumps, which was situated partly in the shoreline just across the A-fiord from Maarmorilik (Figure 1), was later in 1990 removed to the extent possible and the majority of the waste rock, a total of 3.2×10 5 tons were discharged on top of the tailings deposit in the A-fiord.

Methods and instrumentation

Numerous sample collections have been performed in Maarmorilik in the period 1972-2009. The following is a description of the sampling that is basis of the results presented in this paper.

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Sampling of seawater and sediment

Sampling of seawater was done in the A-fiord between 1975 and 2007. All samples were taken in August. Depth-specific samples were taken at 10 m intervals either by a metal-free water sampler or pumped up through a silicone tube. Subsequently, samples were filtered through a 0.45 µm polycarbonate filter into acid-cleaned polyethylene bottles. To preserve samples, 1 ml of Suprapure nitric acid was added to 1 L of seawater immediately after collection. In 2009, seawater was also collected directly from shore at 8 coastal sites (T12SV to L) (Figure 1) and given the same treatment as above. The particulate filtrates (>0.45 µm in size) were kept for subsequent analyses.

Furthermore, sediment was collected from the seafloor at the same locations. In 2005, cores of sediment (0 to ~15 cm) were taken using a HAPS box corer at several locations including Station 12 in Qaamarujuk (Figure 1). The stainless steel HAPS box corer has an inner diameter of 13.5 cm.

The sediment core was cut into slices at 1 cm intervals and sediment that had been in contact with the HAPS was carefully removed prior to sampling. No bioturbation could be observed. Sediment samples were kept in polyethylene bags at freezing temperatures before being sent to Denmark.

Sampling of seaweed

Resident seaweed (Fucus vesiculosus or Fucus distichus) was sampled in the tidal zone directly from shore at the coastal sites (Figure 1) every 1-3 years between 1988 and 2009. Furthermore, to investigate if growth tips of seaweed sampled at the coastal sites were representative to the recent annual metal uptake, some seaweed plants living on small moveable rocks were taken from an unpolluted reference site (Site L) and transplanted to the contaminated sites and collected the following year. A number of 1 to 5 replicate seaweed samples (most often 2) were taken each time.

One sample typically consists of 10 to 15 individual seaweed plants. After sampling, the seaweed was rinsed 3 times in clean water and only the fresh growth tips were kept in order to measure the most recent metal uptake. The growth tips were cut from the rest of the plants using stainless steel scissors and kept in polyethylene bags at freezing temperatures until freeze dried and homogenized in an agate mortar in Denmark prior to chemical analyses.

Sampling of blue mussels

Blue mussels (Mytilus edulis) were collected at the coastal sites at low tide in the period 1984-2009.

To investigate long-and short-term uptake, both resident mussels and mussels transplanted from an unpolluted area (Site L) and collected the following year were sampled. For transplantation,

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6 approximately 50 mussels were placed in two small closed nets connected with a 1 m cord and the cord was squeezed in between a couple of large rocks for subsequent retrieval. Due to the impact of sea ice and wave activity on the shoreline, not all the transplanted mussels were successfully retrieved the following year explaining the lack of data for some years and sites. After collection, the lengths of the mussels were measured in the laboratory and the mussels divided into size classes according to their shell lengths: 4-5 cm, 5-6 cm etc. The mussels were counted, measured more precisely and the mean shell length calculated. One sample generally consists of 20 individual mussels. The adductors of the mussels were cut using a stainless steel scalpel and the shells opened.

The mussels were then allowed to drain for a few minutes before the soft parts were removed from the shells. The soft parts of each size group were pooled into a polyethylene bag and frozen before transported to Denmark. In Denmark, samples were freeze dried and homogenized to powder in an agate mortar prior to chemical analyses. The water content of the sample was determined by weighting the sample before and after freeze drying.

Chemical analyses

Dried sediment subsamples (500 mg) were digested in a mixture of 0.25 ml/0.75 ml/3 ml Suprapure HNO 3 /HCl/HF in Berghof bombs at 120°C for 4 h, then dissolved in milliQ water and the hydrofluoric acid neutralized by boric acid. Dried and homogenized seaweed and blue mussel subsamples (300 mg) were microwave-digested (Anton Paar Multiwave 3000) in Teflon bombs in 4 ml/4 ml Suprapure HNO 3 /milliQ water. Digestion solutions of seaweed and blue mussels were diluted with milliQ water and analysed for Pb and Zn using either flame AAS (Perkin-Elmer 3030), graphite furnace AAS (Perkin-Elmer Zeeman 3030) or ICP-MS (Agilent 7500ce) at NERI, Roskilde, Denmark. Digestion solutions of sediments were diluted with milliQ water and analysed for Pb, Zn, Arsenic (As), cadmium (Cd), mercury (Hg) and aluminum (Al) using ICP-MS. The analytical methods for analyses of sediment and biota have previously been described in detail in [START_REF] Asmund | An update of analytical methods, quality assurance and quality control used in the Greenland AMAP programme 1999-2002[END_REF]. Seawater samples taken before 2007 were analysed for Pb and Zn using Anodic Stripping Voltammetry and a Hg-drop/Hg-film electrode. From 2007, seawater analyses were performed by ICP-MS after pre-concentration and elution of metals on a Chelex-100 filled micro column. In the latter, the pH of seawater samples, standards and reference materials were adjusted to pH 6.3 prior to analyses.

The analytical quality of the above-mentioned analyses was checked by regularly analyzing blanks, duplicates and the certified reference materials DOLT, DORM, TORT, MESS, HISS, PACS, CASS, NASS and SLEW among others. Furthermore, the laboratory at NERI is accredited for

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analyses of elements including Pb, Zn and Cd in biota, sediment and seawater with at precision of 15-20% and participates twice a year in the international QUASIMEME intercalibration program for laboratory performance.

Data analyses

Results from analyses of seawater and seaweed could be used directly for comparison of metal contamination between sites and years without any further data treatment. Data from blue mussels and sediments, however, needed some additional treatment to allow for a comparison.

For blue mussels, the transplantation of mussels has previously been shown to have a negative effect on the condition of the mussels and often results in a reduction of the mussel weight (Riget et al., 1997a). To account for that, the measured metal concentration in µg metal per g dry weight (wt.) mussel tissue was converted into µg metal per mussel by multiplying the measured concentration by the average mussel dry wt. This was done for both transplanted and resident blue mussels.

Furthermore, as the metal contents in blue mussels depend on its length (Riget et al., 1997a) and since the mean lengths of the samples varied, all metal contents in mussels were normalized to the content in a 6 cm blue mussel using the following equation based on data from the same area (Riget et al., 1997a):

54 . 2 _ 6 6 _ _       • = L amount Metal amount Metal L mussel cm
where Metal_amount L is the metal amount in a blue mussel with a mean shell length of L.

Most often, the corrections made to mussel contents were minimal as the preferred size group of 5-6 cm or alternatively 6-7 cm having a mean shell length close to 6 cm could be sampled. In some cases, however, only mussels with a mean length less than 5 cm or more than 7 cm could be found at the site (see Supplementary data section for a list of all blue mussel data).

For sediments, the concentrations of Pb, Zn, As, Cd and Hg are expressed as the ratio of these elements to Al in order to allow for a comparison of metal contents in sediment within the sediment profile at Station 12 and between the coastal sites. The normalization to Al was done to minimize the effect of grain size on the metal concentrations (e.g. [START_REF] Perner | Contamination of arctic fjord sediments by Pb-Zn mining at Maarmorilik in central West Greenland[END_REF] as the sediment samples were not size-fractionated prior to analyses (see Supplementary data section for a list of all sediment data). Time trends of Pb and Zn in seaweed and blue mussels were investigated statistically following the ICES (International Council for the Exploration of the Sea) temporal trend assessment procedure (Nicholson et al., 1998). In short, the log-mean concentration is used as the annual index value. The

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ACCEPTED MANUSCRIPT 8 total variation over time is partitioned into a linear and a non-linear component. Linear regression analysis is applied to describe the linear component and a LOESS smoother (locally weighted quadratic least-squares regression smoothing) with a window width of 7 years is applied to describe the non-linear component. The linear and non-linear components are tested by means of an analysis of variance. The theory behind the use of smoothers in temporal trend analyses is described in detail by [START_REF] Fryer | Using Smoothers for comprehensive assessments of contaminant time series in marine biota[END_REF]. A significance level of 5% was applied.

Results and discussion

Trends in seawater in the A-fiord and Q-fiord

The results from analyses of seawater taken in the A-fiord in autumn during the period are shown in Figure 2. In autumn, the water in the A-fiord is distinctly stratified and a pycnocline is observed at approximately 25 m's depth [START_REF] Johansen | Environmental monitoring at the lead-zinc mine in Maarmorilik[END_REF]. In Figure 2, Pb and Zn concentrations are presented as mean values above and below the pycnocline. During the mining period, bottom water in the A-fiord was heavily contaminated by Pb and Zn with concentrations up to 440 µg L -1 and 791 µg L -1 , respectively. In the surface water, maximum concentrations during the mining period were 30-50 times lower than in the bottom water with Pb and Zn concentration up to 12 µg L -1 and 16 µg L -1 , respectively. The very contaminated bottom water in the A-fiord during the mining period is primarily considered the result of dissolution of Pb and Zn from the tailings after they were discharged but before they settled on the bottom of the A-fiord (Asmund 1992b;[START_REF] Poling | Importance of geochemistry: The Black Angel Lead-Zinc Mine, Greenland[END_REF]. The elevated concentrations of metals in the surface water of the A-fiord are mainly regarded as a consequence of mixing of surface water with metal-rich bottom water during the winter period [START_REF] Johansen | Environmental monitoring at the lead-zinc mine in Maarmorilik[END_REF].

In 1990, when the mine closed and 320,000 tons of waste rock was dumped in the A-fiord, a peak in Pb and Zn concentrations in the surface water and in Zn in the bottom water of the A-fiord was observed. In contrast to Zn, the Pb concentration in the bottom water was lower in 1990 compared to the previous years (Figure 2). The relatively higher content of Zn in seawater in 1990 compared to Pb is considered the result of a higher release of Zn relative to Pb (by a factor of 12) from the waste rock (Asmund, 1992a). After 1990, Pb concentrations decreased abruptly in the A-fiord and during the most recent measurements in 2008, concentrations of 0.086 µg Pb L -1 in surface water and 0.307 µg Pb L -1 in bottom water were measured. Mean Pb concentrations in the bottom water during the period 1995-2008 were reduced to about 1/1000 of those in 1988-89 and Pb levels in the surface water in 2008 were near Pb concentrations in uncontaminated seawater (about a factor of 2 higher than Pb at Site T37; Table 1). Zn levels in the A-fiord have not decreased as much as Pb. In

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2008, the Zn concentration in bottom water was 27 µg L -1 , which is only about 10 times lower than in 1988-89 and 50-100 times higher than in uncontaminated near-shore seawater (Site T37; Table 1). Similarly, the Zn concentration in surface water in the A-fiord was 6.7 µg L -1 in 2008, which is about a factor of 10-50 higher than in uncontaminated near-shore seawater (Table 1). The relative high concentrations of Zn in seawater in the A-fiord indicate that there is still a significant ongoing release of Zn from minerals in the buried tailings and waste rock (presumably ZnS) deposited on the bottom of the A-fiord while a release of Pb from Pb-containing minerals can probably be considered negligible. The observed increase in Zn concentrations in bottom and surface water from 2002 to 2008 (Figure 2) is not readily explained but may reflect a decrease in water exchange between the A-and Q-fiord at the time.

In August 2009, seawater samples were taken directly from shore at several coastal sites in the Qfiord and the metal concentrations are shown in Table 1. At Site T12SV, situated in the former waste rock dump area, near-shore seawater contained high dissolved Pb and Zn concentrations (0.46 and 2.4 µg L -1 , respectively). This is twice the Pb concentration and 1/10 of the Zn concentration at the bottom of the A-fiord. These concentrations are high considering that it is a coastal tidal zone where considerable mixing of seawater takes place. At the remaining sites, Zn concentrations were at approximately the same level and close to uncontaminated seawater (Site T37) while Pb concentrations may be slightly elevated at Site T22 to T36. It has to be noted that the water samples above represent only conditions during the short moment of sampling and may not be representative to conditions during a longer period of time. However, the results indicate that the remains of the former waste rock dump near Site T12SV are still a major source of Pb and Zn contamination in the area.

Trends in sediments in the Q-fiord

A sediment core was taken at Station 12 in the middle of the Q-fiord just outside from Maarmorilik in 2005 and the concentrations of Pb, Zn, As, Cd and Hg normalized to Al in the sediments can be seen in Figure 3. Elevated concentrations of all these elements were observed in sediments that were deposited during the mining period and in the years after but concentrations have decreased in recent years. However, the concentrations of Pb, Zn, As and Hg in the uppermost sediment layer at Station 12 were still elevated in 2005 compared to pre-mining levels. Only Cd had decreased to premining levels. It is interesting to note that Pb, Zn, As and Hg concentrations peaked some years after the mine closure in 1990, which may be the result of a slow long-term sediment transport within the Q-fiord [START_REF] Elberling | Geochemical trends in metalcontaminated fiord sediments near a former lead-zinc mine in West Greenland[END_REF]. The contaminated sediments observed in the middle of

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the Q-fiord is likely to be the result of both a dispersal of fine-grained particles coming from the land-based waste rock dumps as well from metals dissolved from the tailings and waste rock, which were later bound to natural sediment. During the mining period, it was estimated that about equal amounts of both Pb and Zn were released to the Q-fiord from the waste rock dumps (mainly as particles) and as dissolved metals from the tailings, respectively (Asmund, 1992b).

In 2009, near-shore seafloor sediments were sampled together with particulate filtrates (>0.45 µm in size) of seawater from several coastal sites in the Q-fiord (Table 1). The Pb and Zn contents normalized to Al were (except from near-shore sediment at Site T12SV) lower than measured in sediment at Station 12 in the middle of the Q-fiord. The normalized Pb and Zn contents decreased in both near-shore sediments and particulate filtrates with increasing distance to the mine. At Site T12SV and T22, closest to the mine, both near-shore sediments and particulate filtrates contained elevated Pb contents (normalized to Al) compared to the rest of the sites. Elevated Zn contents in near-shore sediments and particulate filtrates were measured in a larger area covering the sites T12SV and T22, but also the more distant sites T25, T29 and T30. The results above indicate that near-shore sediments and particles suspended in the seawater at several of the coastal sites were still metal-contaminated almost two decades after mine closure.

Initial seaweed experiments

The seaweed species Fucus vesiculosus was preferred when sampling seaweed in Maarmorilik.

However, at some sites and during some years it was not possible to find this species and the species Fucus distichus was sampled instead. The two different species look very similar and first it was important to find out whether the metal contents contained in the growth tips of these two species provided similar results. Therefore, both species were sampled at the same sites during some years and the results compared. The results for Pb in growth tips of the two species can be seen in Figure 4a. The results show that there is no systematic difference in metal concentration in growth tips between the two species in Maarmorilik. A similar conclusion was found in a seaweed study from the Nuuk fiord in Southwest Greenland (Riget et al., 1997b).

Next, we investigated whether growth tips of resident seaweed in Maarmorilik represent the uptake of metals within the recent year. To do that, seaweed plants were collected at an uncontaminated site (Site L) and transplanted to various sites in Maarmorilik and collected the following year. The metal concentrations in growth tips of transplanted seaweed were then compared to resident seaweed and the results for Pb can be seen in Figure 4b. Data shows no systematic difference between transplanted and resident seaweed indicating that metal contents in growth tips of resident

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11 seaweed in Maarmorilik can be regarded as an integrated measure of the recent annual uptake of metals. Uptake of metals in seaweed is regarded as a relative measure of the dissolved metal concentration within the surrounding seawater [START_REF] Rainbow | Biomonitoring of heavy metal availability in the marine environment[END_REF][START_REF] Larsen | Lead and zinc in sediments and biota from Maarmorilik, West Greenland: an assessment of the environmental impact of mining wastes on an Arctic fjord system[END_REF]. Consequently, it was concluded that the variations in metal concentration measured in growth tips of resident Fucus vesiculosus or Fucus distichus species at the various sites and years in Maarmorilik could be treated as a relative measure of the recent annual variation in dissolved metal concentrations at the sampling sites.

Metal contamination of seaweed in Maarmorilik and temporal and spatial trends

Metal contamination of seaweed related to mining in Maarmorilik is shown in Table 2. Here Pb, Zn, As, Cd and Hg concentrations in seaweed growth tips at an uncontaminated site (Site L) versus at a Seaweed at Site T12SV contained the most Pb and Zn with concentrations up 84 µg g -1 and 363 µg g -1 dry wt., respectively. The high concentrations measured in 1990 are considered due to the removal of the old shore-line waste rock dump near Site T12SV and subsequently the discharge of waste rock into the A-fiord releasing Pb and Zn in the process. An additional factor was the processing of old oxidized ore and the following release of tailings with a high soluble Pb content during the last months before mine closure. It is believed that the discharge of waste rock is the main explanation for the high Zn concentrations in 1990 and the processing of old oxidized ore is the main reason for the high Pb concentrations in 1990 (Asmund, 1992a;Asmund, 1992b).

Since 1990, seaweeds at contaminated sites in Maarmorilik area have shown decreasing trends in Pb and Zn concentrations (Table 3). This applies to all sites (a total of 21) within a distance of 20 km from the mine (Schiedek et al., 2009). At most sites the trends were significant at a 5 % significance level and trends were both indicated by log-linear trends (exponential decreasing) and non-linear trends (Table 3). The dominant trend, however, was log-linear when considering all seaweed sites near Maarmorilik (Schiedek et al., 2009). The decreasing trends in Pb and Zn further away from Maarmorilik (Site L was not measured). In contrast, however, elevated Zn concentrations in seaweed from Site T22 to T36 were not reflected in measureable elevated dissolved Zn concentrations in seawater at those sites. The observed difference is likely due to the fact that metal concentrations in seaweed growth tips represent a complex time-integrated measure of the dissolved metal concentrations in the seawater during the whole growing season while the seawater samples represent only conditions during the short period of sampling.

Metal contamination of blue mussels in Maarmorilik and temporal and spatial trends

The uptake of Pb, Zn, As, Cd and Hg in blue mussels near Maarmorilik is exemplified in Table 2. Pb and 1,190 µg g -1 dry wt. Zn. This corresponds to 2,530 µg Pb and 970 µg Zn in a blue mussel with a shell length of 6 cm. According to previous studies, suspension feeders such as blue mussels take up metals bound to particulate matter and food particles as well as metals in dissolved form within the seawater [START_REF] Rainbow | Biomonitoring of heavy metal availability in the marine environment[END_REF][START_REF] Larsen | Lead and zinc in sediments and biota from Maarmorilik, West Greenland: an assessment of the environmental impact of mining wastes on an Arctic fjord system[END_REF]. Consequently, the high Pb and Zn concentrations in 1989 and 1990 are considered the result of both a higher dispersion of dissolved and particle-bound Pb and Zn at that time as a consequence of removal and subsequently disposal of 3). This applies to all blue mussel sites (a total of 19) within a distance of 35 km from the mine (Schiedek et al., 2009). For transplanted mussels, decreases in Pb contents (and Zn at the site nearest to the mine) were also observed after 1990. At the more distant sites, Zn contamination in mussels was very low or negligible during the period when transplantation of mussels was conducted. Time trends for resident blue mussels after 1990 were significant at all sites at a 5 % significance level and trends were predominately log-linear (exponential decreasing) (Table 3). The decrease in Pb and Zn in resident and transplanted blue mussels is considered the result of a decrease in dissolution of Pb and Zn from tailings and waste rock in the A-fiord (now buried under a layer of natural sediment) as well as a decrease in outwash of dissolved and particle-bound Pb and Zn from the remaining land-based waste rock dumps near Maarmorilik. For resident mussels, the decrease in Pb and Zn is also influence by a disappearance of old mussels in the size group preferred and the emergence of new generations of mussels that have been exposed to lower concentrations.

This table shows element concentrations in blue mussels

In 2009, Pb contents in transplanted blue mussels within an area of 12 km from Maarmorilik were still elevated by factors of 2-25 after one year of transplantation (Figure 8). In contrast, relative Zn levels were much lower and were less than a factor of 2 higher than the initial content for all sites (Figure 8). The limited contamination by Zn in blue mussels is in contrast to seaweed, which contains elevated concentrations of Zn in a larger area around Maarmorilik. The observed differences between Zn in blue mussels and seaweed may be explained by differences in sources of metal uptake between the two species i.e. metal uptake by blue mussels may be more controlled by particle-bound metals than dissolved metals. An alternative explanation is that blue mussels may have a greater capacity to regulate the internal accumulation of Zn compared to seaweed, Zn being an essential nutritional element to biota in contrast to Pb.

Transplanted versus resident blue mussels

Transplanted blue mussels were used for monitoring recent metal contamination at the coastal sites in addition to sampling of resident blue mussels. A previous study from Maarmorilik has shown that once metals such as Pb and Zn are taken up in blue mussels, these are released at a very slow rate should the environmental conditions improve (Riget et al, 1997a). Furthermore, blue mussels grow relatively slowly in arctic waters and a blue mussel with the preferred shell length of 6 cm is

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ACCEPTED MANUSCRIPT 14 likely to be approximately 12 years old [START_REF] Theisen | The growth of Mytilus edulis L. (bivalvia) from Disko and Thule district, Greenland[END_REF]. Consequently, resident blue mussels did not seem adequate for monitoring annual variations in seawater contamination in a situation with decreasing contamination. Thus, blue mussels were transplanted to selected sites from an uncontaminated site (Site L) and collected the following year. The differences between the Pb and Zn contents in transplanted versus resident blue mussels at typical sites can be seen in Figure 7. The Pb and Zn contents were consistently lower in transplanted mussels after one year of transplantation compared to resident mussels. The largest differences between transplanted and resident mussels were observed just after the mining period when the change in contamination rate was highest and the concentrations approached each other towards the end of the monitoring period. Also, it is notable that the Pb content in transplanted mussels at Site T17A (Figure 7) in 1985 were about 4 times higher than in 1992 while the resident mussels contained approximately the same amount.

The differences between Pb and Zn in transplanted versus resident blue mussels after 1990 show that Pb and Zn are released from the resident mussels at a very slow rate. It also shows that transplantation of blue mussels is an adequate way to monitor short-term (annual) changes in seawater metal contamination in contrast to resident mussels. This is at least the case for monitoring metals such as Pb and Zn in arctic blue mussels in a situation where the contamination is decreasing.

Contamination sources, processes and "lessons learned" in Maarmorilik

During the mining period from 1973 to 1990, the major inputs of mine-related metals to the marine environment in Maarmorilik came from disposal of mine tailings and later waste rock into the small partly-enclosed A-fiord (Asmund, 1992b). The A-fiord is separated from the outer Q-fiord by a shallow submarine sill and an almost stagnant water mass is created below the sill during summer. This presumably makes the A-fiord an ideal trap for suspended particles such as mine tailings, which at the time when mining was commenced was thought to contain insoluble metal compounds.

However, as studies later revealed, the tailings contained heavy metals including Pb, Zn and Cd bound in chemical forms that were indeed soluble in the ambient seawater. The main minerals of the tailings were calcite, dolomite and pyrite but it was indicated that most of the Zn and Cd were contained in the mineral sphalerite (ZnS). Pb was mainly contained in minerals other than sulphides, probably carbonates or sulphates [START_REF] Loring | Heavy metal contamination of a Greenland Fiord system by mine wastes[END_REF]Asmund, 1992b). Studies showed that metals dissolved during the dilution of the tailings with the seawater and the simultaneous decrease of pH from approximately 10 to 8 (Asmund, 1992b). Seasonal mixing of the water column in the A-fiord was later found to take place during winter as a result of freezing of the

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15 surface water and an inflow of water between the A-fiord and Q-fiord [START_REF] Møller | Hydrodynamics of an arctic fiord[END_REF]. The complete vertical mixing of water and increased water exchange between the two fiords increased the dispersion of dissolved metals as well as fine suspended tailings particles from the A-fiord to the Qfiord. Also, the mixing of water led to increased oxidation-and dissolution-potentials of the tailings particles in suspension as well as the tailings and waste rock situated on the bottom of the A-fiord [START_REF] Elberling | Geochemical trends in metalcontaminated fiord sediments near a former lead-zinc mine in West Greenland[END_REF]. During the mining period, the main process responsible for the input of dissolved metals to the marine environment in Maarmorilik was considered the dissolution of mine tailings. The dissolution process occurred mainly after discharge but before settling of the tailings close to the outlet point (Asmund, 1992b). This was shown by an abrupt decrease in Pb concentrations in the bottom water of the A-fiord after the mine closed down in 1990 and the last produced tailings had settled. While the main input of dissolved metals came from the tailings, the waste rock dumps were significant sources of contamination of the intertidal zones (Asmund, 1992b).

After 1990, the Pb concentrations decreased abruptly in the bottom water of the A-fiord while the Zn concentrations did not decrease to the same extent. During the last measurements in 2008, Zn levels were still 50-100 times higher than in uncontaminated seawater. This indicates that considerable amounts of Zn were and still are released from the buried mine waste in the A-fiord while the dissolution of Pb are probably negligible. In contrast, the abrupt decrease in Pb concentration in water within the A-fiord was not reflected in a similar abrupt decrease in Pb contamination in seaweed and blue mussels. This indicates that the main sources of Pb contamination of the marine environment in Maarmorilik have likely shifted from the tailings disposed into the A-fiord to outwash of dissolved and particle-bound Pb from the land-based waste rock dumps as well as from land-deposited dust that was spread during the mining period. Given the time trends for Pb and Zn in seaweed and blue mussels, it seems likely that elevated levels of Pb and Zn in seaweed and blue mussels will still be measured for at least 50 more years near Maarmorilik, even if no new mining activity is commenced.

In the Maarmorilik case, poor environmental studies and decision-making led to significant metal contamination of the marine environment near Maarmorilik. Several lessons can be learned from that, including those listed below. First of all, if mine waste is going to be disposed into the sea, very detailed studies of how the waste will react in that environment has to be made prior to disposal because once it is disposed it is nearly impossible to retrieve. Studies prior to disposal must include representative waste material (tailings/waste rock) obtained from a pilot project prior to the mine operation subjected to the conditions in which it is going to be disposed (using representative

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ACCEPTED MANUSCRIPT 16 seawater). More rough tests, such as leach tests using acetic acid will also provide valuable results.

The effects of different redox conditions on the chemical behavior of the mine waste have to be included in the tests as well as a thorough understanding of the hydrological conditions all year round at the disposal-site as these may effect the mobilization and dispersion of contaminants.

Bioaccumulation studies are useful tools to reveal how different organisms react to the given exposure. During mining activities, attention must be given to minimize the generation and dispersal of dust as dust particles may remain in the terrestrial environment for several decades after the initial deposition. Finally, chemically reactive waste rock must be kept away from the tidal zone and if at all possible transported to a confined area as part of the mining process and subsequently covered e.g. with chemically inert rock material such as the marble/dolomite found in Maarmorilik.

The remains of a former land-based waste rock dump is still a major contaminant source two decades after mine closure, which shows that once large amounts of waste rock are dumped in an environmentally unsafe manner, it is extremely difficult to remediate afterwards probably due to finer grained particles remaining on site.

Conclusion

This study assessed the present environmental state adjacent to the former Pb-Zn mine in Maarmorilik, West Greenland. Despite the two decades that have passed since mine closure; seawater, sediment and biota, including seaweed and blue mussels are still contaminated with Pb and Zn related to the past mining operation. This reveals a slow recovery of this arctic fiord system to mining pollution. Specifically, it was shown that 12 kilometers from the mine, seaweed and blue mussels still contain elevated concentrations of Pb (and in seaweed also Zn). During the mining period, the primary source of contamination was considered the dissolution of metals from mine tailings that were discharged into the small fiord Affarlikassaa near Maarmorilik. Later, following the settling of tailings on the bottom of Affarlikassaa and the burial of these with natural sediment, the main sources of Pb contamination of the marine environment in Maarmorilik have likely changed. Remaining land-deposited waste rock, especially the residues of an old large-scale waste rock dump as well as outwash Pb-contaminated dust generated during the mining period are now considered the dominant sources of Pb contamination in Maarmorilik. The contamination by Zn, however, still seems to be considerably influenced by a release of Zn from the waste materials at the bottom of Affarlikassaa. Table 1. Pb and Zn concentrations in seawater filtered to <0.45 µm in size; in the remaining particulate filtrate; and in near-shore sediment sampled at various sites near Maarmorilik in 2009. In the filtrate and sediment, the Pb/Al and Zn/Al ratios are also shown in order to normalize the Pb and Zn contents against variations in grain size. Table 2. Pb, Zn, As, Cd and Hg concentrations (µg g -1 dry wt.) in growth tips of seaweed and in blue mussel tissue in mussels with a shell length of 5-6 cm sampled in 2009. Results are given for seaweed samples collected at an uncontaminated reference site (Site L) and at a high-impact site (Site T12SW) and for blue mussels from Site L before and after one year of transplantation at Site T12SV.

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Tables

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ACCEPTED MANUSCRIPT 20 Table 3. Results of the temporal trend analyses of Pb and Zn in seaweed and in resident and transplanted blue mussels at three sites in Maarmorilik with increasing distances to the mine/former waste rock dump area sampled during the period after mining (1991)(1992)(1993)(1994)(1995)(1996)(1997)(1998)(1999)(2000)(2001)(2002)(2003)(2004)(2005)(2006)(2007)(2008)(2009). Site T12SW (a) is situated in the former waste rock dump area; Site T17A (b) and Site T36 (c) are located about 2 and 12 km away from Maarmorilik, respectively (Figure 1). Significance at the 5% level is shown by "sign" and non-significance by "-" for both the log-linear trend and the non-linear trend components. The results of the trend analyses can be interpreted as follows: 1) Both log-linear and non-linear trend not significant = no temporal trend; 2) Log-linear trend significant, non-linear trend not significant = log-linear trend (exponential trend); 3) Both log-linear trend and non-linear trend significant = non-linear trend; 4) Log-linear trend not significant, non-linear trend significant = nonlinear trend. Furthermore, the overall annual change in percentage during the period is given. Transplanted mussels were collected at an unpolluted reference site (Site L), transplanted in nets to the three sites and sampled the year after. Since the mussel sizes varied, the metal contents have been normalized to the contents in a 6 cm blue mussel (see Data analyses section in text).

Figures

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Research highlights for the article: "Long-term response of an arctic fiord system to submarine disposal of mine waste: The environment at Maarmorilik, West Greenland two decades after mine closure."

• A case-study of an arctic fiord systems long-term response to mining pollution.

• Two decades after mine closure the fiord is still contaminated with lead and zinc.

• Metals are still elevated in seaweed and blue mussels twelve km from the mine.

• Zinc is still released from marine-disposed mine tailings while lead is probably not.

• Transplantation of blue mussels provides a way of monitoring annual contamination.

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Table S-1. Resident and transplanted blue mussels from selected sites in Maarmorilik sampled from 1985 to 2009. Transplanted mussels were collected at an unpolluted reference site (Site L), transplanted in nets, and collected at the sampling sites one year after. When different size groups exist, mussels with a length of 5-6 cm is preferred, thereafter 6-7 cm, and finally 4-5 cm.

  dumps was removed to the extent possible and the waste rock disposed on top of the tailings in the A-fiord as part of the mine closure plan in 1990.

  Figure 5 for three contaminated sites with increasing distances to the mine. At all sites, the highest Pb and Zn concentrations in seaweed during the period 1988-2009 were measured in 1990.

  contamination in seaweed from 1990 to 2009 are considered the result of a decrease in dissolution M A N U S C R I P T A C C E P T E D ACCEPTED MANUSCRIPT of Pb and Zn from the tailings and waste rock discharged in the A-fiord and from the remaining waste rock left on the steep mountain sides near Maarmorilik.In 2009, when the most recent sample was taken, Pb and Zn concentrations in seaweed within an area of 12 km from Maarmorilik were still elevated by factors of 7-30 for Pb and 4-16 for Zn (Figure6). The elevated levels of Pb and Zn in seaweed correspond partly with the Pb and Zn levels measured in seawater at the same sites in 2009 (Table2). The elevation of Pb and Zn in seaweed at Site T12SV in 2009 was consistent with elevated dissolved Pb and Zn in seawater at measured at Site T12SV. Similarly, elevated Pb concentrations in seaweed from Site T22 to T36 were consistent with higher dissolved Pb in seawater sampled in 2009 at those sampling sites compared to Site T37

  transplanted from an uncontaminated reference site (Site L) in 2008 to the most contaminated site in Maarmorilik (Site T12SV) and collected in 2009. The results show that only Pb and Zn were elevated after one year of transplantation (by a factor of 25 for Pb and ~2 for Zn) indicating that Pb and Zn are the elements of main importance with respect to metal contamination of blue mussels in Maarmorilik. Pb and Zn concentrations in resident and transplanted blue mussels from three contaminated sites in Maarmorilik during the period 1985-2009 are shown in Figure 7. At all sites, the highest Pb and Zn concentrations in mussels during the period were measured in 1989 and 1990. Resident blue mussels at Site T12SV contained the most Pb and Zn with concentrations up to 2,650 µg g -1 dry wt.

  rock dump into the A-fiord as well as disposal of mine tailings with a higher soluble Pb content during the last months of mine operation.After 1990, resident blue mussels at contaminated sites in Maarmorilik area have shown decreasing trends in Pb and Zn concentrations (Table

Figure 1 .

 1 Figure 1. Map of Maarmorilik and the surrounding area with the location of the sampling sites.

Figure 2 .

 2 Figure 2. Mean Pb and Zn concentrations in surface water (0-20 m depth) (a) and in bottom water (30-60 m depth) (b) in Affarlikassaa during autumn (august) from 1975 to 2008.

Figure 3 .

 3 Figure 3. Variations of Pb, Zn, As, Cd and Hg in sediment sampled in the Qaamarujuk fiord in 2005 at Station 12, just outside Maarmorilik. The concentrations have been normalized to Al in order to compensate for variations in grain size. The dotted lines indicate the approximate position of layers sedimented in 1973 and 1990 based on sedimentation rates and 210 Pb dating of sediments at the same site performed by[START_REF] Elberling | Geochemical trends in metalcontaminated fiord sediments near a former lead-zinc mine in West Greenland[END_REF].

Figure 4 .

 4 Figure 4. Pb concentrations in growth tips from the seaweed species Fucus distichus versus Fucus vesiculosus (a); and Pb concentrations in growth tips from resident seaweed versus growth tips from seaweed transplanted from an unpolluted area (Site L) and collected at the site one year after (b).

Figure 5 .Figure 6 .

 56 Figure 5. Pb and Zn concentrations in growth tips from seaweed (Fucus distichus or Focus vesiculosus) at three sites in Maarmorilik with increasing distances to the mine/former waste rock dump sampled from 1988 to 2009.

Figure 7 .

 7 Figure 7. Pb and Zn contents in resident and transplanted blue mussels at three sites in Maarmorilik with increasing distances to the mine/former waste rock dump sampled from 1985 to 2009.

Figure 8 .

 8 Figure 8. Pb and Zn contents in blue mussels transplanted from Site L in 2008 and collected at five sites near Maarmorilik in 2009. The distances from the former waste rock dump area are shown in parentheses. The metal contents have been normalized to the contents in a 6 cm blue mussel (see Data analyses section in text).

  Figure 1. Søndergaard et al., 2011

Table S -

 S 2. Metal concentrations in sediment, in mg kg -1 , measured in a sediment profile taken at Station 12 in Qaamarujuk in 2005.

						ACCEPTED MANUSCRIPT ACCEPTED MANUSCRIPT ACCEPTED MANUSCRIPT ACCEPTED MANUSCRIPT ACCEPTED MANUSCRIPT
	T25 T5 T5 Site	Nat Nat Nat	2000 1994 1989 Dissolved (<0.45 µm) 4-5 20 8-9 -6-7 -	4.6 8.5 6.5	2.4 16.6 7.7 Particulate (>0.45 µm) 0.3 14 29 2.3 14 566 1.2 15 658	224 436 430	9 1312 772	72 1010 504 Near-shore sediment 19 145 542 417 630 411
	T30 T36 T12SW Nat Nat Nat T17A Nat T12SW Nat Depth (cm) T17A Nat 0.5	2000 2000 1994 1994 1989 Zn Pb 1989 236 (µg L -1 )	4-5 4-5 7-8 7-8 7-9 5-6	20 20 ---Pb Zn -173 (µg L -1 )	4.6 4.5 7.5 7.5 8.0 Pb Cd 5.5 0.18 (µg g -1 ) 2.3 2.5 11.5 12.9 14.3 5.4	0.3 0.4 1.7 1.7 2.0 Zn 0.8 0.021 Hg (µg g -1 )	13 16 15 13 14 Pb/Al 48 18 984 480 2647 14 334 (10 -4 )	233 165 469 529 676 Zn/Al As 322 19 (10 -4 )	15 7 1659 810 5259 Pb Al 256 39616 (µg g -1 )	73 65 791 892 1343 247 (µg g -1 ) 30 14 942 459 2533 Zn 320	147 132 449 506 Pb/Al 647 308 (10 -4 )	Zn/Al (10 -4 )
	Station T22 T25 T30 T36 T5 T12SW Trans Resident (Nat) or transplant ed (Trans) Nat Nat Nat Nat Trans T17A Trans T22 Trans T36 Trans T22 Trans T36 Trans T5 Nat T12SW Nat T17A Nat T25 Nat T30 Nat T36 Nat T5 Nat T12SW Nat T17A Nat T22 Nat T25 Nat T30 Nat T36 Nat T5 Nat T12SW Nat T17A Nat T22 Nat T25 Nat T30 Nat T36 Nat T5 Trans T17A Trans T22 Trans T25 Trans T36 Trans T5 Nat T12SW Nat T17A Nat T22 Nat T25 Nat T30 Nat T36 Nat T5 Nat T12SW Nat T17A Nat T22 Nat T36 Trans T36 Nat T30 Trans T30 Nat T25 Trans T25 Nat T22 Trans T22 Nat T17A Trans T17A Nat T12SW Trans T12SW Nat T5 Trans T5 Nat T36 Nat T36 Nat T25 Nat T30 Nat T36 Nat T5 Trans T12SW Trans T17A Trans T22 Trans T25 Trans T30 Trans T36 Trans T5 Nat T12SW Nat T17A Nat T22 Nat T25 Nat T30 Nat T17A Nat T22 Nat T25 Nat T30 Nat T36 Nat T17A Trans T22 Trans T25 Trans T30 Trans T36 Trans T5 Nat T12SW Nat T17A Nat T22 Nat T25 Nat T30 Nat T5 Trans T12SW Trans T17A Trans T22 Trans T25 Trans T30 Trans T36 Trans T5 Nat T12SW Nat T17A Nat T22 Nat T25 Nat T30 Nat T36 Nat T5 Trans T12SW Trans T17A Trans T22 Trans T25 Trans T30 Trans T36 Trans T5 Nat T12SW Nat T17A Nat T22 Nat T22 Nat T25 Nat T30 Nat T36 Nat T5 Trans T17A Trans T22 Trans T25 Trans T30 Trans T36 Trans T5 Nat T12SW Nat T17A Nat T22 Nat T25 Nat T30 Nat T36 Nat T5 Trans T17A Trans T22 Trans T25 Trans T30 Trans T36 Trans T5 Nat T12SW Nat T22 Nat T25 Nat T30 Nat T36 Nat T5 Nat T12SW Nat T17A Nat T22 Nat T25 Nat T30 Nat T36 Nat T17A Nat T17A Nat T17A Nat T17A Trans * See the "Data analyses" section in the text for explanation. Year Size group (cm) Number of individuals 2009 5-6 16 2009 5-6 20 2009 5-6 18 2009 5-6 17 2009 5-6 11 2009 5-6 8 2009 5-6 5 2009 5-6 6 2009 5-6 10 2008 5-7 14 2008 5-6 19 2008 5-6 20 2008 5-6 18 2008 5-6 20 2008 5-6 20 2008 5-6 20 2008 5-6 20 2007 5-6 20 2007 5-6 20 2007 5-6 20 2007 5-6 20 2007 5-6 20 2007 5-6 20 2007 5-6 20 2005 5-6 20 2005 5-6 20 2005 5-6 20 2005 5-6 20 2005 5-6 20 2005 5-6 20 2005 5-6 20 2002 5-6 36 2002 5-6 20 2002 4-6 5 2002 5-7 20 2002 5-7 19 2002 5-6 20 2002 5-6 20 2002 5-6 20 2002 5-6 20 2002 5-6 20 2002 5-6 20 2002 5-6 20 2000 4-5 20 2000 4-5 20 2000 4-5 20 2000 4-5 20 1995 -7 1990 5-6 -1995 -7 1990 6-7 -1995 -6 1990 6-7 -1995 -7 1990 7-8 -1995 -7 1990 6-7 -1995 -7 1990 6-7 -1995 -5 1990 6-7 -1995 -18 1991 5-6 20 A C C E P T E D Mean length (cm) Mean ind. wet wt. (g) Mean ind. dry wt. (g) Dry wt. (%) 5.5 4.6 0.8 17 5.5 4.5 0.7 15 5.5 4.4 0.7 15 5.5 3.5 0.6 16 5.4 4.5 0.5 12 5.5 5.4 0.8 14 5.4 5.2 0.7 14 5.3 4.0 0.6 14 5.4 3.7 0.5 13 6.0 5.3 0.6 12 5.3 3.6 0.5 14 5.5 8.0 1.1 13 5.6 4.5 0.6 14 5.4 4.3 0.6 14 5.5 4.2 0.6 14 5.5 3.7 0.5 14 5.7 3.9 0.3 9 5.5 6.0 1.0 16 5.5 5.2 0.6 12 5.5 4.0 0.6 16 5.5 3.2 0.5 16 5.5 5.9 0.9 15 5.5 3.6 0.4 12 5.5 3.8 0.4 12 5.5 5.7 1.2 21 5.6 5.5 1.0 18 5.5 3.9 0.6 15 5.5 4.4 0.7 17 5.5 4.7 0.7 16 5.6 4.3 0.6 15 5.5 4.4 0.7 15 5.8 5.0 0.6 11 5.6 4.3 0.3 7 4.9 3.2 0.4 11 6.0 6.5 0.7 11 6.1 5.3 0.6 11 5.5 5.2 0.9 16 5.5 4.6 0.7 15 5.4 4.2 0.7 16 5.5 4.8 0.8 16 5.5 4.7 0.7 15 5.5 4.6 0.6 13 5.5 4.6 0.7 14 4.5 2.4 0.3 13 4.6 2.8 0.4 15 4.5 2.5 0.4 15 4.5 2.3 0.3 14 6.8 11.3 2.2 19 5.5 5.9 0.8 13 6.6 7.1 1.3 18 6.5 8.1 1.2 14 6.7 12.2 2.4 20 6.5 8.2 1.2 14 6.7 10.7 2.1 20 7.5 10.9 1.7 15 6.6 7.7 1.2 16 6.5 8.5 1.2 14 6.6 10.4 2.0 20 6.5 8.1 1.2 15 6.5 13.1 2.9 22 6.5 7.4 1.2 16 7.8 14.5 3.2 22 5.5 4.8 1.0 20 1996 -20 7.7 9.9 1.6 16 1996 -20 7.3 7.7 1.1 15 1996 -19 6.8 7.9 1.1 13 1996 -20 4.9 3.7 0.6 16 1996 -21 5.1 3.4 0.5 14 1996 -18 4.9 2.6 0.3 13 1996 -18 5.9 3.2 0.5 15 1996 -11 5.6 4.8 0.7 15 1996 -10 5.1 3.1 0.4 12 1996 -18 4.8 2.4 0.3 13 1995 -11 7.6 16.9 3.8 23 1995 -20 7.8 17.4 3.4 20 1995 -20 6.9 10.6 2.2 20 1995 -17 7.5 15.0 3.5 23 1995 -20 7.3 13.0 2.6 20 1995 -20 7.3 10.9 2.0 18 A C C E P T E D 1992 5-6 -5.5 5.6 0.8 14 1992 7-8 -7.5 10.6 1.7 16 1992 6-7 -6.5 7.2 1.2 16 1992 6-7 -6.5 8.2 1.2 14 1992 5-6 -5.5 4.8 0.8 16 1992 7.0-7.4 -7.2 11.1 1.5 14 1992 7.0-7.4 -7.2 13.2 1.5 11 1992 7.0-7.4 -7.2 10.9 1.5 14 1992 7.5-7.9 -7.7 14.7 1.8 12 1992 7.0-7.4 -7.2 11.7 1.5 13 1991 7-8 19 7.5 12.7 2.0 15 1991 6-7 20 6.5 8.4 1.3 16 1991 5-6 20 5.5 4.1 0.6 14 1991 6-7 17 6.5 8.0 1.3 16 1991 6-7 19 6.5 7.1 1.2 16 1991 6-7 19 6.5 6.4 1.1 17 A C C E P T E D A C C E P T E D A C C E P T E D M Pb (μg/g dry wt.) 32 12 15 20 15 70 18 16 9 14 7 42 227 37 18 29 20 21 155 31 34 21 26 16 13 139 26 24 12 17 14 26 33 66 12 10 29 113 40 37 20 30 15 127 191 55 51 4 233 8 460 5 332 11 567 15 509 82 2031 10 845 72 176 178 341 69 33 272 37 39 15 25 11 194 655 166 180 116 114 301 441 237 198 139 28 28 13 19 15 718 972 310 406 276 335 A N U S C R I P T Zn (μg/g dry wt.) Pb pr. mussel (μg) Zn pr. mussel (μg) Pb in 6 cm mussel (μg)* Zn in 6 cm musling (μg)* 129 25 99 32 126 109 8 75 10 92 128 10 86 13 105 136 11 75 14 92 274 8 143 10 185 245 53 185 66 230 209 13 156 17 201 298 9 169 12 229 229 4 108 6 140 176 8 108 8 108 205 4 103 5 139 163 45 176 57 224 474 144 301 174 363 117 22 70 29 91 123 10 70 13 88 112 16 60 20 77 123 7 41 8 47 120 20 116 25 145 386 98 245 120 300 135 19 84 24 105 171 17 88 21 109 212 18 184 23 231 174 12 78 14 150 7 66 9 83 102 15 119 19 149 272 138 271 168 328 133 15 76 19 95 130 17 95 21 117 140 9 103 11 128 126 11 79 13 95 163 9 107 12 135 282 15 162 16 176 442 11 141 13 169 393 23 139 39 234 220 8 151 9 153 267 6 155 6 149 163 24 139 30 173 350 80 248 100 311 187 27 125 36 197 28 149 34 201 15 146 19 225 18 137 23 151 10 98 12 458 40 145 85 584 79 242 159 224 21 85 42 229 16 75 34 156 95 8 207 6 151 428 179 328 223 410 173 132 11 168 8 132 503 539 590 440 481 486 109 11 262 9 198 553 389 648 318 529 306 99 23 206 18 156 576 956 971 543 551 123 157 18 193 14 152 637 597 747 487 609 171 213 168 435 132 341 1012 2381 1187 1943 968 184 87 29 255 24 208 603 991 707 809 577 185 139 230 446 118 229 319 169 306 210 381 167 253 287 407 152 216 352 382 394 232 240 226 73 240 53 174 202 20 121 33 203 563 131 270 197 408 283 12 93 20 156 264 18 124 19 129 217 11 154 13 184 273 9 98 13 149 227 3 68 6 120 224 739 853 405 468 290 2247 995 1154 511 231 357 497 250 348 208 632 730 358 414 205 300 531 183 323 187 223 367 136 223 349 231 268 288 334 445 744 750 422 426 429 278 503 227 410 353 232 414 189 338 229 107 176 133 219 180 42 274 26 172 262 43 398 27 251 171 20 260 12 164 210 35 379 18 201 190 23 289 14 182 382 1404 747 796 424 404 1295 538 1057 439 317 182 186 227 232 373 525 483 429 394 375 322 438 263 357 465 363 504 296 411 97 1999 4-6 17 5.0 3.8 0.5 14 16 222 8 113 13 180 1999 4-6 16 5.1 3.5 0.4 11 202 458 79 179 122 277 1999 5-7 20 5.0 2.7 0.3 13 29 249 10 84 16 134 1999 5-6 21 5.3 4.0 0.5 12 17 235 8 109 11 147 1999 5-6 22 5.8 4.7 0.5 11 11 207 6 106 6 115 1999 5-6 28 5.4 4.3 0.5 12 12 214 7 114 9 149 1999 5-7 13 5.9 5.3 0.6 11 12 242 7 145 7 149 1998 -30 3.9 2.3 0.4 17 23 114 9 45 27 137 1998 -15 8.2 11.7 1.4 12 1173 598 1681 857 765 390 1998 -17 7.0 8.6 1.0 11 224 248 220 243 147 163 1998 -30 3.3 1.1 0.2 16 38 180 6 30 29 139 1998 -17 7.9 9.8 1.1 11 169 319 186 352 92 174 1998 -15 8.1 10.7 1.1 11 252 281 284 317 135 150 1998 -18 7.1 8.5 0.9 10 117 216 104 193 67 124 1997 -16 5.4 6.0 0.7 12 23 201 17 148 22 194 1997 -18 4.8 3.8 0.4 11 154 471 63 194 111 340 1997 -16 5.4 5.1 0.6 12 26 226 16 142 21 188 1997 -19 5.6 4.7 0.7 14 18 221 12 151 14 179 1997 -11 5.5 5.9 0.7 11 9 181 6 118 8 146 1997 -20 5.3 3.6 0.4 11 11 200 5 82 6 114 1997 -20 5.7 5.9 0.6 10 7 218 4 128 5 149 1996 -20 7.8 11.7 2.1 18 329 329 684 684 351 351 1996 -20 7.1 8.3 1.3 16 1126 526 1486 694 969 453 1996 -19 6.8 8.3 1.3 15 152 279 195 357 142 260 1996 -19 7.9 12.7 1.9 15 304 322 578 612 287 304 M A N U S C R I P T 1994 7-8 -7.5 9.9 1.7 17 390 432 658 729 373 413 1994 7-8 -7.5 11.2 1.7 15 252 374 425 631 241 358 1994 7-8 -7.5 13.1 1.7 13 292 212 492 358 279 203 1994 7-8 -7.5 11.8 1.7 14 159 311 268 524 152 298 1994 7.5-7.9 -7.7 12.0 1.8 15 28 130 51 234 27 124 1994 7.0-7.4 -7.2 13.2 1.5 12 18 176 28 268 17 168 1994 6-7 -6.5 8.5 1.2 14 26 170 30 199 25 163 1994 6-7 -6.5 9.0 1.2 13 14 183 16 215 13 175 1994 6-7 -6.5 10.7 1.2 11 19 164 22 192 18 157 1994 6-7 -6.5 6.6 1.2 18 7 131 8 154 7 125 1993 7-8 20 7.5 12.8 2.1 16 594 378 1253 798 711 452 1993 6-7 18 6.5 9.9 1.2 13 1155 472 1427 583 1164 476 1993 6-7 20 6.5 6.8 0.9 14 425 346 397 323 324 264 1993 8-9 18 8.5 15.5 2.2 14 369 324 804 706 332 291 1993 6-7 19 6.5 8.2 0.9 12 313 347 297 330 243 269 1993 6-7 20 6.5 7.3 1.0 14 261 337 266 344 217 280 1993 5-6 15 5.5 5.7 0.8 14 109 350 87 280 109 349 1993 7.2-7.4 3 7.2 12.9 1.9 15 18 165 35 317 22 199 1993 7.2-7.4 7 7.2 11.1 1.5 14 18 176 27 268 17 168 1993 7.2-7.4 7 7.2 9.7 1.2 12 18 205 22 240 14 151 1993 7.2-7.4 7 7.2 12.4 1.7 14 14 120 25 208 16 131 1993 7.2-7.4 7 7.2 10.4 1.1 11 17 251 18 275 12 173 1993 7.2-7.4 7 7.2 10.2 1.2 11 13 203 15 236 9 149 1992 7-8 -7.5 11.9 1.7 14 733 426 1236 718 701 408 1992 6-7 -6.5 8.3 1.2 14 1124 467 1318 548 1075 447 M A N U S C R I P T 1989 6-7 -6.5 7.4 1.2 16 628 494 736 579 601 473 1989 6-7 -6.5 8.0 1.2 15 341 449 400 526 326 430 1989 6-7 -6.5 8.6 1.2 14 322 407 378 477 308 389 1989 5-6 -5.5 4.7 0.8 16 115 240 88 184 110 230 1988 6-7 -6.5 8.0 1.2 15 566 405 664 475 542 388 1988 6-7 -6.5 7.4 1.2 16 1536 558 1801 654 1470 534 1988 5-6 -5.5 4.9 0.8 16 330 392 253 301 316 375 1988 6-7 -6.5 6.9 1.2 17 498 489 584 573 476 468 1988 6-7 -6.5 6.8 1.2 17 264 433 310 508 253 414 1988 5-6 -5.5 5.4 0.8 14 314 436 241 334 300 417 1988 5-6 -5.5 4.8 0.8 16 168 371 129 285 161 355 1987 5-6 17 5.5 4.8 0.8 16 292 282 222 214 276 266 1986 5-6 19 5.4 4.9 0.6 12 239 276 143 166 188 217 1985 5-6 12 5.3 4.9 0.8 17 285 354 231 287 316 393 1985 -5 7.2 10.3 1.3 12 131 193 168 247 107 158 M A N U S C R I P T 1.5 262 186 0.30 0.023 21 T12SV 0.46 2.44 75 11 31 218 213 436 69 40799 2.5 298 187 0.75 0.022 18 T22 0.07 0.47 14 1 21 291 6 42 12 40939 3.5 265 94 0.87 0.018 13 T25 0.11 0.78 34 2 5 96 4 38 2 45965 4.5 165 45 0.43 0.014 11 T29 0.11 0.77 29 1 5 107 8 35 1 45205 5.5 143 34 0.24 0.010 10 T30 0.12 0.52 31 2 5 70 15 55 3 42719 6.5 144 32 0.28 0.010 10 45642 7.5 143 29 0.26 0.010 10 44482 8.5 131 28 0.23 0.011 10 43892 9.5 132 28 0.20 0.009 10 45072 10.5 130 26 0.21 0.010 10 45592 11.5 125 25 0.21 0.007 10 43737 12.5 126 26 0.21 0.009 10 45407 13.5 124 25 0.23 0.009 10 44745 M T36 0.06 0.13 26 6 2 7 9 20 2 T37 0.04 0.63 25 1 3 53 7 27 2 L ND a ND 15 3 2 11 14 13 3 A N U S C R I P T a ND = Not determined	139 80 22 6 13 3 5 2
									6

Table 1 .

 1 Søndergaard et al., 2011.

	M A N U S C R I P T

Table 2 .

 2 Søndergaard et al., 2011.

	M A N U S C R I P T

Table 3 .

 3 Søndergaard et al., 2011