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Marine Environmental Research 71 (2011) 331e341
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Marine Environmental Research
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Long-term response of an arctic fiord system to leadezinc mining and submarinedisposal of mine waste (Maarmorilik, West Greenland)
Jens Søndergaard*, Gert Asmund, Poul Johansen, Frank RigétDepartment of Arctic Environment, National Environmental Research Institute, Aarhus University, Frederiksborgvej 399, DK-4000 Roskilde, Denmark
a r t i c l e i n f o
Article history:Received 10 November 2010Received in revised form15 March 2011Accepted 16 March 2011
Keywords:MiningSubmarine tailings disposalPollution monitoringHeavy metalsLeadZincMytilus edulisFucus vesiculosusGreenlandArctic
* Corresponding author. Tel.: þ45 46 30 19 66.E-mail address: [email protected] (J. Søndergaard).
0141-1136/$ e see front matter � 2011 Elsevier Ltd. Adoi:10.1016/j.marenvres.2011.03.001
a b s t r a c t
Contamination by lead (Pb) and zinc (Zn) was studied in seawater, sediments, seaweeds and blue musselsnear the former Black Angel PbeZn Mine in Maarmorilik, West Greenland. The mine operated during theperiod 1973e90 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 dischargearea were measured up to 440 and 790 mg L�1, respectively. Pb in fiord sediments, seaweeds and bluemussels just outside the discharge area were measured in concentrations up to 190, 84 and 2650 and Znup to 300, 360 and 1190 mg g�1 dry wt., respectively. Within the discharge area, seawater metalconcentrations (especially Pb) decreased abruptly after mine closure. Metals concentrations in sedimentsand biota, however, decreased more slowly and two decades after mine closure seaweeds and bluemussels were still contaminated 12 km from the mine.
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1. Introduction
The Black Angel LeadeZinc Mine in Maarmorilik in WestGreenland (Fig. 1) is an important example in mining history wheredischarge of waste materials into the sea caused a significantrelease and dispersion of metals within a larger fiord system(Asmund, 1992a; Larsen et al., 2001; Elberling et al., 2002). TheBlack Angel Mine operated between 1973 and 1990 and during thisperiod, so-called mine tailings from ore treatment were dischargedinto a small partly-enclosed fiord Affarlikassaa (named the A-fiord).In the years after tailings disposal began, elevated concentrations ofmetals including lead (Pb) and zinc (Zn) were found in seawater,sediments and biota, not only in the A-fiord but also in a larger areacovering the outer fiord Qaamarujuk (Q-fiord). Later, detailedgeochemical studies revealed that the tailings contained significantamounts of Pb and Zn containing minerals that were soluble in theambient seawater (Poling and Ellis, 1995). Furthermore, seasonaldestratification of the water body in the A-fiord was found to takeplace during winter/spring. This allowed for a complete mixing of
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surface water with metal-rich bottom water and subsequentlya dispersal of contaminants over the sill between the A-fiord andthe Q-fiord (Poling and Ellis, 1995). Other sources of contaminationin Maarmorilik included dust from the ore-milling process andwaste rock dumps established on land below the mine entrancescontaining rocks with elevated levels of Pb and Zn. The largest andmost contaminating of the waste rock dumps was partly removed(it was not possible to remove all material) and the waste rockdisposed on top of the tailings in the A-fiord as part of the mineclosure plan in 1990.
In 2010, a re-opening of the mine was planned within the nextcoming years and this study seeks to evaluate the state of theenvironment in the A- and Q-fiords 20 years after mine closure.This is important not only as an environmental “baseline study” forthe re-opening of the mine but also as a study of long-termrecovery to an arctic fiord system subjected to mining pollution,which may also provide a valuable reference for future cases ofsubmarine tailings disposal. The study is based on analyses ofseawater, sediments and the marine organisms e seaweeds (Fucusvesiculosus/Fucus distichus) and blue mussels (Mytilus edulis). Theserelatively sedentary organisms concentrate metals, reflect differentbioavailable metal-sources within the seawater and have been usedas biological monitors of heavy metal pollution in a number of
Fig. 1. Map of Maarmorilik and the surrounding area with the location of the samplingsites, the tailings disposal area and the site where a large waste rock dump was partlyremoved in 1990.
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341332
previous studies, including the so-called Mussel Watch Programs(Rainbow, 1995; Cairrao et al., 2007; O’Connor and Lauenstein,2006; Torres et al., 2008). Data include samples taken during themining period (before 1990) and until 2009 and provide a detailedconsistent record of high-resolution data from environmentalmonitoring and research in the area.
2. Site description
Maarmorilik is situated adjacent to the two fiords, Affarlikassaaand Qaamarujuk (A- and Q-fiords), in the inner part of the Uum-mannaq fiord complex in West Greenland (71�070 N; 51�150 W)(Fig. 1). The closest community is Ukkussissat, 25 km to the westand Uummannaq, the main settlement in the area, is situated80 km south of Maarmorilik. The climate is arctic with maximumsummer temperatures around 10 �C and minimum wintertemperatures below �30 �C. Winds are dominated by strongeasterly winds coming from the Greenland Ice Sheet in the bottomof the Q-fiord, which result in sparse precipitation in the area(<100 mm a year).
The A-fiord covers an area of 2 km2 and has an average waterdepth of 30 m. The fiord is partly separated from the outer Q-fiordby a sill at 23 m depth. The A- and Q-fiords can be described asstratified estuaries during summer. The stratification is caused bymelt water from the rivers that flow on top of the heavier seawatercreating a distinct and roughly stagnant water mass in the bottomof the Q-fiord and A-fiord. The annual fresh water input to theQ-fiord equals approximately 90 � 106 m3 from the river in thebottom of the A-fiord and 22 � 106 m3 from the Wegener River inthe inner part of the Q-fiord (Møller, 1984). In almost every winter,however, a complete vertical mixing of the stratified water bodiestakes place due to freezing of surface water and water inflow overthe sill between the A-fiord and Q-fiord (Møller, 1984). The annualseawater temperatures in the fiords range between�2 and 5 �C, thesalinity between 31 and 34 psu and the pH between 7.8 and 8.1. Thetidal range in the fiords is around 1 m and sea ice, 0.5e1 m thick,usually develops in October and covers the fiords until May.
The geological Maarmorilik Formation includes manycarbonate-hosted PbeZn 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 CentralWest Greenland (Escher and Pulvartaft, 1976). The name “BlackAngel” refers to a pelite outcrop forming a dark angel-like figure
high on a 1100 m marble cliff face above the A-fiord. The massiveores, up to 30 m thick, consist of galena (PbS), sphalerite (ZnS) andpyrite (FeS2) with accessory ore minerals such as pyrrhotite, chal-copyrite, tennantite and arsenopyrite. Prior to mining, the BlackAngel deposit consisted of ten major ore bodies with a total of13.6 � 106 tons containing 4.0% Pb, 12.3% Zn and 29 ppm Ag. Out ofthese, 11.2 � 106 tons were mined in the period 1973e90 by thecompany Greenex A/S (Thomassen, 2003).
During the mining process, ore material was transported fromthe mine entrances at 600 m altitude by means of cable cars acrossthe A-fiord to a floatation plant in Maarmorilik. Here, concentrateswere produced, loaded onto ships and transported to smelters inEurope. For ore treatment, conventional rod/ball milling wasapplied to liberate Pb and Zn. This was followed by froth floatationto produce separate Pb and Zn concentrates plus a waste product,tailings (Poling and Ellis, 1995). Permission for submarine disposalof tailings was given to the mining company and tailings weresubsequently discharged as a suspension in seawater and leddirectly into the A-fiord at around 30 m depth. In the 1970s, theannual tailings inputs from the flotation mill to the A-fiord were4.5 � 105 t, which increased to 6 � 105 t in the 1980s. During themining period it is estimated that more than 8 � 106 t of minetailings containing more than 2.2 � 104 t Pb and 5 � 104 t Zn weredischarged into the A-fiord (Elberling et al., 2002). As part of themine exploitation, several waste rock dumps were establishedbelow the mine entrances in the Black Angel Mountain. The mostpolluting of these waste rock dumps, which was situated partly inthe shoreline just across the A-fiord from Maarmorilik (Fig. 1), waslater in 1990 removed to the extent possible and themajority of thewaste rock, a total of 3.2 � 105 tons were discharged on top of thetailings deposit in the A-fiord (Asmund, 1992a).
3. Methods and instrumentation
Numerous sample collections were performed inMaarmorilik inthe period 1972e2009. The following is a description of thesampling that is basis of the results presented in this paper.
3.1. Sampling of seawater and sediments
Sampling of seawater was done in the A-fiord between 1975 and2007. All samples were taken in August. Depth-specific sampleswere taken at 10 m intervals either by a metal-free water sampleror pumped up through a silicone tube. Subsequently, samples werefiltered through a 0.45 mm polycarbonate filter into acid-cleanedpolyethylene bottles. To preserve samples, 1 ml of Suprapure nitricacid was added to 1 L of seawater immediately after collection. In2009, seawater was also collected directly from shore at 8 coastalsites (T12SV to L) (Fig. 1) and given the same treatment as above.The particulate filtrates (>0.45 mm in size) were kept for subse-quent analyses. Furthermore, sediments were collected from theseafloor at the same locations. In 2005, cores of sediments (0 tow15 cm) were taken using a HAPS box corer at several locationsincluding Station 12 in Qaamarujuk (Fig.1). The stainless steel HAPSbox corer has an inner diameter of 13.5 cm. The sediment coreswere cut into slices at 1 cm intervals and sediment that had been incontact with the HAPS was carefully removed prior to sampling. Nobioturbation could be observed. Sediment samples were kept inpolyethylene bags at freezing temperatures before being sent toDenmark.
3.2. Sampling of seaweeds
Resident seaweed (F. vesiculosus or F. distichus) was sampled inthe tidal zone directly from shore at the coastal sites (Fig. 1) every
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341 333
1e3 years between 1988 and 2009. Furthermore, to investigate ifgrowth tips of seaweed sampled at the coastal sites wererepresentative to the recent annual metal uptake, some seaweedplants living on small moveable rocks were taken from anunpolluted reference site (Site L) and transplanted to thecontaminated sites and collected the following year. A number of1e5 replicate seaweed samples (most often 2) were taken eachtime. One sample typically consists of 10e15 individual seaweedplants. After sampling, the seaweed was rinsed 3 times in cleanwater and only the fresh growth tips were kept in order tomeasure the most recent metal uptake. The growth tips were cutfrom the rest of the plants using stainless steel scissors and keptin polyethylene bags at freezing temperatures until freeze driedand homogenized in an agate mortar in Denmark prior tochemical analyses.
3.3. Sampling of blue mussels
Blue mussels (M. edulis) were collected at the coastal sites atlow tide in the period 1984e2009. To investigate long- and short-term uptake, both resident mussels and mussels transplantedfrom an unpolluted area (Site L) and collected the following yearwere sampled. For transplantation, approximately 50 musselswere placed in two small closed nets connected with a 1 m cordand the cord was squeezed in between a couple of large rocks forsubsequent retrieval. Due to the impact of sea ice and waveactivity on the shoreline, not all the transplanted mussels weresuccessfully retrieved the following year explaining the lack ofdata for some years and sites. After collection, the lengths of themussels were measured in the laboratory and the mussels dividedinto size classes according to their shell lengths: 4e5 cm, 5e6 cmetc. The mussels were counted, measured more precisely and themean shell length calculated. One sample generally consists of 20individual mussels. The adductors of the mussels were cut usinga stainless steel scalpel and the shells opened. The mussels werethen allowed to drain for a few minutes before the soft parts wereremoved from the shells. The soft parts of each size groupwere pooled into a polyethylene bag and frozen before trans-ported to Denmark. In Denmark, samples were freeze dried andhomogenized to powder in an agate mortar prior to chemicalanalyses. The water content of the sample was determined byweighting the sample before and after freeze drying.
3.4. Chemical analyses
Dried sediment subsamples (500 mg) were digested ina mixture of 0.25 ml/0.75 ml/3 ml Suprapure HNO3/HCl/HF inBerghof bombs at 120 �C for 4 h, then dissolved in milliQ water andthe hydrofluoric acid neutralized by boric acid. Dried and homog-enized seaweed and blue mussel subsamples (300 mg) weremicrowave-digested (Anton Paar Multiwave 3000) in Teflon bombsin 4 ml/4 ml Suprapure HNO3/milliQ water. Digestion solutions ofseaweed and blue mussels were diluted with milliQ water andanalyzed for Pb and Zn using either flame AAS (PerkineElmer3030), graphite furnace AAS (PerkineElmer Zeeman 3030) or ICP-MS (Agilent 7500ce) at NERI, Roskilde, Denmark. Digestion solu-tions of sediments were diluted with milliQ water and analyzed forPb, Zn, Arsenic (As), cadmium (Cd), mercury (Hg) and aluminum(Al) using ICP-MS. The analytical methods for analyses of sedimentand biota have previously been described in detail in Asmund et al.(2004). Seawater samples taken before 2007 were analyzed for Pband Zn using Anodic Stripping Voltammetry and a Hg-drop/Hg-filmelectrode. From 2007, seawater analyses were performed by ICP-MSafter pre-concentration and elution of metals on a Chelex-100 filled
micro column. In the latter, the pH of seawater samples, standardsand reference materials were adjusted to pH 6.3 prior to analyses.
The analytical quality of the above-mentioned analyses waschecked by regularly analyzing blanks, duplicates and the certifiedreference materials DOLT, DORM, TORT, MESS, HISS, PACS, CASS,NASS and SLEW among others. Furthermore, the laboratory at NERIis accredited for analyses of elements including Pb, Zn and Cd inbiota, sediment and seawater with at precision of 15e20% andparticipates twice a year in the international QUASIMEME inter-calibration program for laboratory performance.
3.5. Data analyses
Results from analyses of seawater and seaweed could be useddirectly for comparison of metal contamination between sites andyears without further data treatment. Data from blue mussels andsediments, however, needed some additional treatment to allowfor a comparison.
For blue mussels, the transplantation of mussels has previouslybeen shown to have a negative effect on the condition of themussels and often results in a reduction of the mussel weight (Rigetet al., 1997a). To account for that, themeasuredmetal concentrationin mg metal per g dry weight (wt.) mussel tissue was converted intomg metal per mussel by multiplying the measured concentration bythe averagemussel dry wt. This was done for both transplanted andresident blue mussels. Furthermore, as the metal contents in bluemussels depend on its length (Riget et al., 1997a) and since themean lengths of the samples varied, all metal contents in musselswere normalized to the content in a 6 cm blue mussel using thefollowing equation based on data from the same area (Riget et al.,1997a):
Metaleamount6cmemussel ¼ MetaleamountL,�6L
�2:54
where Metal_amountL is the metal amount in a blue mussel witha mean shell length of L.
Most often, the corrections made to mussel contents wereminimal as the preferred size group of 5e6 cm or alternatively6e7 cm having a mean shell length close to 6 cm could be sampled.In some cases, however, only mussels with a mean length less than5 cm or more than 7 cm could be found at the site (seeSupplementary data section for a list of all blue mussel data).
For sediments, the concentrations of Pb, Zn, As, Cd and Hg wereexpressed as the ratio of these elements to Al in order to allow fora comparison of metal contents in sediment within the sedimentprofile at Station 12 and between the coastal sites. The normali-zation to Al was done to minimize the effect of grain size on themetal concentrations (e.g. Perner et al., 2010) as the sedimentsamples were not size-fractionated prior to analyses (seeSupplementary data section for a list of all sediment data).
Time trends of Pb and Zn in seaweed and blue mussels wereinvestigated statistically following the ICES (International Councilfor the Exploration of the Sea) temporal trend assessment proce-dure (Nicholsen et al., 1995). In short, the log-mean concentrationwas used as the annual index value. The total variation over timewas partitioned into a linear and a non-linear component. Linearregression analysis was applied to describe the linear componentand a LOESS smoother (locally weighted quadratic least-squaresregression smoothing) with a windowwidth of 7 years was appliedto describe the non-linear component. The linear and non-linearcomponents were tested by means of an analysis of variance. Asignificance level of 5% was applied. The theory behind the use ofsmoothers in temporal trend analyses is described in detail by Fryerand Nicholson (1999).
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341334
4. Results and discussion
4.1. Trends in seawater in the A-fiord and Q-fiord
The results from analyses of seawater taken in the A-fiord inautumn during the period 1975e2008 are shown in Fig. 2. Inautumn, the water in the A-fiord is distinctly stratified and a pyc-nocline is normally observed at approximately 25 m’s depth(Johansen et al., 2008). In Fig. 2, Pb and Zn concentrations are pre-sented as mean values above and below the pycnocline. During themining period, bottom water in the A-fiord was heavily contami-nated by Pb and Zn with concentrations up to 440 mg L�1 and790 mg L�1, respectively. In the surface water, maximum concen-trations during the mining period were 30e50 times lower than inthe bottomwater with Pb and Zn concentration up to 12 mg L�1 and16 mg L�1, respectively. The very contaminated bottom water in theA-fiord during the mining period is primarily considered the resultof dissolution of Pb and Zn from the tailings after they were dis-charged but before they settled on the bottom of the A-fiord(Asmund, 1992b; Poling and Ellis, 1995). The elevated concentra-tions of metals in the surface water of the A-fiord are mainlyregarded as a consequence of mixing of surface water with metal-rich bottom water during the winter period (Johansen et al., 2008).
In 1990, when the mine closed and 3.2 � 105 tons of waste rockwas dumped in the A-fiord, a peak in Pb and Zn concentrations inthe surface water and in Zn in the bottomwater of the A-fiord wasobserved. In contrast to Zn, the Pb concentration in the bottomwater was lower in 1990 compared to the previous years (Fig. 2).
0
2
4
6
8
10
12
14
1970 1975 1980 1985 1990
) L / g µ ( b P
0
100
200
300
400
500
1970 1975 1980 1985 1990
Yea
) L / g µ ( b P
a
b
Mining period
Fig. 2. Mean Pb and Zn concentrations in surface water (0e20 m depth) (a) and in bottom wto 2008.
The relatively higher content of Zn in seawater in 1990 compared toPb is considered the result of a higher release of Zn relative to Pb (bya factor of 12) from the waste rock (Asmund, 1992a). After 1990, Pbconcentrations decreased abruptly in the A-fiord and during themost recent measurements in 2008, concentrations of0.086 mg Pb L�1 in surface water and 0.310 mg Pb L�1 in bottomwater were measured. Mean Pb concentrations in the bottomwaterduring the period 1995e2008 were reduced to about 1/1000 ofthose in 1988e89 and Pb levels in the surface water in 2008 werenear Pb concentrations in uncontaminated seawater (about a factorof 2 higher than Pb at Site T37; Table 1). Zn levels in the A-fiord havenot decreased as much as Pb. In 2008, the Zn concentration inbottom water was 27 mg L�1, which is only about 10 times lowerthan in 1988e89 and 50e100 times higher than in uncontaminatednear-shore seawater (Site T37; Table 1). Similarly, the Zn concen-tration in surface water in the A-fiordwas 6.7 mg L�1 in 2008, whichis about a factor of 10e50 higher than in uncontaminated near-shore seawater (Table 1). The relative high concentrations of Zn inseawater in the A-fiord indicate that there is still a significantongoing release of Zn from minerals in the buried tailings andwaste rock (presumably ZnS) deposited on the bottom of the A-fiord while a release of Pb from Pb-containing minerals can prob-ably be considered negligible. The observed increase in Znconcentrations in bottom and surface water from 2002 to 2008(Fig. 2) is not readily explained but may reflect a decrease in waterexchange between the A- and Q-fiord at the time.
In August 2009, seawater samples were taken directly fromshore at several coastal sites in the Q-fiord and the metal
1995 2000 2005 2010 0
2
4
6
8
10
12
14
16
18
) L / g µ ( n Z
Pb (0-20 m) Zn (0-20 m)
1995 2000 2005 2010
r
0
200
400
600
800
) L / g µ ( n Z
Pb (30-60 m)
Zn (30-60 m)
ater (30e60 m depth) (b) in the Affarlikassaa fiord during autumn (August ) from 1975
0 0 2 0 4 0 6 0 8 0
Element/Al
Table 1Pb and Zn concentrations in seawater filtered to<0.45 mm in size; in the remaining particulate filtrate; and in near-shore sediment sampled at various sites near Maarmorilikin 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.
Site Dissolved (<0.45 mm) Particulate (>0.45 mm) Near-shore sediment
Pb (mg L�1) Zn (mg L�1) Pb (mg g�1) Zn (mg g�1) Pb/Al (10�4) Zn/Al (10�4) Pb (mg g�1) Zn (mg g�1) Pb/Al (10�4) Zn/Al (10�4)
T12SV 0.46 2.44 11 75 31 218 213 436 69 139T22 0.07 0.47 1 14 21 291 6 42 12 80T25 0.11 0.78 2 34 5 96 4 38 2 22T29 0.11 0.77 1 29 5 107 8 35 1 6T30 0.12 0.52 2 31 5 70 15 55 3 13T36 0.06 0.13 6 26 2 7 9 20 2 3T37 0.04 0.63 1 25 3 53 7 27 2 5L NDa ND 3 15 2 11 14 13 3 2
Source: Søndergaard et al., 2011.a ND ¼ Not determined.
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341 335
concentrations are shown in Table 1. At Site T12SV, situated in theformer waste rock dump area, near-shore seawater contained highdissolved Pb and Zn concentrations (0.46 and 2.4 mg L�1, respec-tively). This is twice the Pb concentration and 1/10 of the Znconcentration at the bottom of the A-fiord. These concentrationsare high considering that it is a coastal tidal zone where consid-erable mixing of seawater takes place. At the remaining sites, Znconcentrations were at approximately the same level and close touncontaminated seawater (Site T37) while Pb concentrations maybe slightly elevated at Site T22 to T36. It has to be noted that thewater samples above represent only conditions during the shortmoment of sampling and may not be representative to conditionsduring a longer period of time. However, the results indicate thatthe remains of the former waste rock dump near Site T12SV are stilla major source of Pb and Zn contamination in the area.
2
4
6
8
10
12
14
Zn/Al(x10 -4 )
Pb/Al(x10 -4 )
Hg/Al (x10 -8 )
As/Al (x10 -5 )
Cd/Al (x10 -6 )
) m
c (
h t p e D
1973
1990
Fig. 3. Pb, Zn, As, Cd and Hg in sediment sampled in the Qaamarujuk fiord just outsideMaarmorilik in 2005. Concentrations have been normalized to Al in order tocompensate for variations in grain size. The dotted lines indicate the approximateposition of layers sedimented in 1973 and 1990 based on sedimentation rates and 210Pbdating of sediments at the same site performed by Elberling et al. (2002).
4.2. 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 concentrations ofPb, Zn, As, Cd andHg normalized to Al in the sediments are shown inFig. 3. Elevated concentrations of all these elements were observedin sediments that were deposited during the mining period and inthe years after but concentrations have decreased in recent years.However, the concentrations of Pb, Zn, As and Hg in the uppermostsediment layer at Station 12were still elevated in 2005 compared topre-mining levels. Only Cd had decreased to pre-mining levels. Thehighest concentrations of Pb, Zn, As, Cd and Hg in the sedimentswere 190, 300, 21, 0.87 and 0.023 mg kg�1, respectively. It is inter-esting to note that Pb, Zn, As and Hg concentrations peaked someyears after the mine closure in 1990, which may be the result ofa slow long-term sediment transport within the Q-fiord (Elberlinget al., 2002). The contaminated sediments observed in the middleof 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 aswell frommetals dissolved from the tailings and waste rock, whichwere later bound to natural sediment. During the mining period, itwas estimated that about equal amounts of both Pb and Zn werereleased to the Q-fiord from the waste rock dumps (mainly asparticles) and as dissolved metals from the tailings, respectively(Asmund, 1992b).
In 2009, near-shore seafloor sediments were sampled togetherwith particulate filtrates (>0.45 mm in size) of seawater fromseveral coastal sites in the Q-fiord (Table 1). The Pb and Zn contentsnormalized to Al were (except from near-shore sediment at SiteT12SV) lower than measured in sediment at Station 12 in themiddle of the Q-fiord. The normalized Pb and Zn contentsdecreased in both near-shore sediments and particulate filtrates
with increasing distance to the mine. At Site T12SV and T22, closestto the mine, both near-shore sediments and particulate filtratescontained Pb contents (normalized to Al) compared to the rest ofthe sites. Elevated Zn contents in near-shore sediments andparticulate filtrates were measured in a larger area covering the
0
5
10
15
20
25
0 5 10 15 20 25
Transplanted seaweed (µg Pb g -1 dry wt.)
g b P
g µ ( d e e
w
a e s t n e d i s e R
1 -
) . t w
y r d
0
20
40
60
80
100
0 20 40 60 80 100
Fucus distichus (µg Pb g -1 dry wt.)
s
u
s
o
l u
c
i s
e
v
s
u
c
u
F
g b P
g µ ( 1 -
) . t w
y r d
a b 1:1 1:1
Fig. 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 versusgrowth tips from seaweed transplanted from an unpolluted area (Site L) and collected at the site one year after (b).
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341336
sites T12SV and T22, but also the more distant sites T25, T29 andT30. The results above indicate that near-shore sediments andparticles suspended in the seawater at several of the coastal siteswere still metal-contaminated almost two decades after mineclosure.
4.3. Initial seaweed experiments
The seaweed species F. vesiculosuswas preferredwhen samplingseaweed in Maarmorilik. However, at some sites and during someyears it was not possible to find this species and the speciesF. distichuswas sampled instead. The two different species look verysimilar and first it was important to find out whether the metalcontents contained in the growth tips of these two species providedsimilar results. Therefore, both species were sampled at the samesites during some years and the results compared. The results for Pbin growth tips of the two species can be seen in Fig. 4a. The resultsshow that there is no systematic difference in metal concentrationin growth tips between the two species in Maarmorilik. A similarconclusion was found in a seaweed study from the Nuuk fiord inSouthwest Greenland (Riget et al., 1997b).
Next, we investigated whether growth tips of resident seaweedin Maarmorilik represent the uptake of metals within the recentyear. To do that, seaweed plants were collected at an uncontami-nated site (Site L) and transplanted to various sites in Maarmorilikandcollected the followingyear. Themetal concentrations ingrowthtips of transplanted seaweed were then compared to residentseaweed and the results for Pb can be seen in Fig. 4b. Data shows no
Table 2Pb, Zn, As, Cd and Hg concentrations (mg g�1 dry wt.) in growth tips of seaweed andin blue mussel tissue in mussels with a shell length of 5e6 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 T12SV) and for blue mussels from Site L beforeand after one year of transplantation at Site T12SV. Concentrations were measuredin one subsample taken from a homogenized bulksample containing growth tipsfrom 10-15 seaweed plants and mussel soft parts from typically 20 individual bluemussels, respectively.
Element Seaweed Blue mussels
Site L Site T12SV Beforetransplantation(Site L)
After one-yeartransplantation atSite T12SV
Pb 0.11 3.0 2.9 70Zn 8.3 130 146 245As 43 72 16 17Cd 0.81 1.1 3.3 2.8Hg 0.070 0.017 0.12 0.11
Source: Søndergaard et al., 2011.
systematic difference between transplanted and resident seaweedindicating thatmetal contents in growth tips of resident seaweed inMaarmorilik can be regarded as an integratedmeasure of the recentannual uptake of metals. Uptake of metals in seaweed is regarded asa relative measure of the dissolved metal concentration within thesurrounding seawater (Rainbow, 1995; Larsen et al., 2001). Conse-quently, it was concluded that the variations inmetal concentrationmeasured in growth tips of resident F. vesiculosus or F. distichusspecies at the various sites andyears inMaarmorilik couldbe treatedas a relative measure of the recent annual variation in dissolvedmetal concentrations at the sampling sites.
4.4. Metal contamination of seaweed in Maarmorilik and temporaland spatial trends
Metal contamination of seaweed related to mining in Maar-morilik is exemplified in Table 2. Here, Pb, Zn, As, Cd and Hgconcentrations measured in seaweed growth tips from an uncon-taminated site (Site L) is shown versus seaweed from a high-impactsite (Site T12SV) sampled in 2009. Of the elements analyzed, Pb andZn were the most elevated with factors of 15e30 times higherconcentrations measured at Site T12SV, indicating that Pb and Znare the elements of main importance and justifying the focus on Pband Zn in the following.
Pb and Zn concentrations in seaweed during the period1988e2009 in Maarmorilik are shown in Fig. 5 for three contami-nated sites with increasing distances to the mine. At all sites, thehighest Pb and Zn concentrations in seaweed during the periodwere measured in 1990. Seaweed at Site T12SV contained the mostPb and Zn with concentrations up to 84 mg g�1 and 360 mg g�1 drywt., respectively. The high concentrations measured in 1990 areconsidered due to removal of the old shoreline waste rock dumpnear Site T12SV and subsequently the discharge of waste rock intothe A-fiord releasing Pb and Zn in the process. An additional factorwas the processing of old oxidized ore and the following release oftailings with a high soluble Pb content during the last monthsbeforemine closure. It is believed that the discharge of waste rock isthe main explanation for the high Zn concentrations in 1990 andthe processing of old oxidized ore is themain reason for the high Pbconcentrations in 1990 (Asmund, 1992a, 1992b).
From 1991 to 2009, seaweeds at contaminated sites in Maar-morilik showed decreasing trends in Pb and Zn concentrations(Table 3). Decreases in Pb and Zn in seaweeds were observed at allsites (a total of 21) included in the regular environmental moni-toring program in Maarmorilik within a distance of 20 km from themine (Schiedek et al., 2009). At most sites, trends were significantat a 5% significance level and trends were both indicated by log-
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Site T12SV (<1 km from mine)
Site 17A (2 km from mine)
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Fig. 5. Mean Pb and Zn concentrations in growth tips from seaweed (Fucus distichus or Fucus vesiculosus) at three sites in Maarmorilik with increasing distances to the mine sampledfrom 1988 to 2009.
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341 337
linear trends (exponential decreasing) and non-linear trends(Table 3). The dominant trend, however, was log-linear whenconsidering all seaweed sites near Maarmorilik (Schiedek et al.,2009). The decreasing trends in Pb and Zn contamination inseaweed from 1991 to 2009 are considered the result of a decreasein dissolution of Pb and Zn from the tailings and waste rock dis-charged in the A-fiord and from the remaining waste rock left onthe steep mountain sides near Maarmorilik.
In 2009, when the most recent samples were taken, Pb and Znconcentrations in seaweed within an area of 12 km from Maar-morilik were still a factor of 7e30 higher for Pb and a factor of 4e16higher for Zn compared to seaweed at Site L, 35 km away (Fig. 6).The higher levels of Pb and Zn in seaweed correspond partly withthe Pb and Zn levels measured in seawater at the same sites in 2009(Table 2). The highest concentrations of Pb and Zn in seaweedas measured at Site T12SV were consistent with the highest
Table 3Results of the temporal trend analyses of Pb and Zn in seaweed and in resident andtransplanted blue mussels at three sites in Maarmorilik with increasing distances tothe mine/former waste rock dump area sampled during the period after mining(1991e2009). Site T12SV (a) is situated in the former waste rock dump area; SiteT17A (b) and Site T36 (c) are located about 2 and 12 km away from Maarmorilik,respectively (Fig. 1). Significance at the 5% level is shown by “sign” and non-significance bye 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 andnon-linear trend not significant¼ no temporal trend; 2) Log-linear trend significant,non-linear trend not significant ¼ log-linear trend (exponential trend); 3) Both log-linear trend and non-linear trend significant ¼ non-linear trend; 4) Log-linear trendnot significant, non-linear trend significant ¼ non-linear trend. Furthermore, theoverall annual change in percentage during the period is given.
Site Element Number ofsamples
Log-lineartrend
Non-lineartrend
Annualchange (%)
SeaweedT12SV Pb 13 sign e �8.1T12SV Zn 13 e e �2.8T17A Pb 12 sign e �5.4T17A Zn 12 sign sign �1.1T36 Pb 13 e e �1.0T36 Zn 13 sign sign �3.6Blue mussels, residentT12SV Pb 12 sign sign �15.4T12SV Zn 12 sign e �2.4T17A Pb 12 sign e �18.2T17A Zn 12 sign e �8.1T36 Pb 13 sign e �18.2T36 Zn 13 sign e �8.6Blue mussels, transplantedT12SV Pb 5 e e �5.9T12SV Zn 5 e e �3.4T17A Pb 9 e e �1.7T17A Zn 9 sign e þ0.8T36 Pb 9 sign sign �3.4T36 Zn 9 e e �0.4
Source: Søndergaard et al., 2011.
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341338
concentrations of dissolved Pb and Zn in seawater. Similarly, highPb concentrations in seaweed from Site T22 to T36 were consistentwith higher dissolved Pb in seawater sampled at those samplingsites compared to Site T37 further away from Maarmorilik (Site Lwas not measured). In contrast, however, high Zn concentrations inseaweed from Site T22 to T36 relative to Site L were not reflected inmeasureable elevated dissolved Zn concentrations in seawater atthose sites. The observed difference is likely due to the fact thatmetal concentrations in seaweed growth tips represent a complextime-integrated measure of the dissolved metal concentrations in
0
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Pb Zn
Fig. 6. Pb and Zn concentrations in growth tips from seaweed (Fucus distichus or Fucusvesiculosus) at five sites in Maarmorilik in 2009. The distances from the mine are shownin parentheses. Each bar represent the concentration measured in one subsample takenfrom a homogenized bulksample with growth tips from 10-15 individual seaweedplants.
the seawater during the whole growing season while the seawatersamples represent only conditions during the short period ofsampling.
4.5. Metal contamination of blue mussels in Maarmorilik andtemporal and spatial trends
The uptake of Pb, Zn, As, Cd and Hg in blue mussels nearMaarmorilik is exemplified in Table 2. The table shows elementconcentrations in blue mussels transplanted from an uncontami-nated reference site (Site L) in 2008 to the most contaminated sitein Maarmorilik (Site T12SV) and collected in 2009. The resultsindicate that only Pb and Zn were elevated after one year oftransplantation (by a factor of 25 for Pb andw2 for Zn) and that Pband Zn are the elements of main importance with respect to metalcontamination of blue mussels in Maarmorilik.
Pb and Zn concentrations in resident and transplanted bluemussels from three contaminated sites in Maarmorilik during theperiod 1985e2009 are shown in Fig. 7. At all three sites, the highestPb and Zn concentrations in mussels during the period weremeasured in 1989 and 1990. Resident blue mussels at Site T12SVcontained the most Pb and Zn with concentrations up to2650 mg g�1 dry wt. Pb and 1190 mg g�1 dry wt. Zn. This correspondsto 2530 mg Pb and 970 mg Zn in a blue mussel with a shell length of6 cm. According to previous studies, suspension feeders such asblue mussels take up metals bound to particulate matter and foodparticles as well as metals in dissolved form within the seawater(Rainbow, 1995; Larsen et al., 2001). Consequently, the high Pb andZn concentrations in 1989 and 1990 are considered the result ofboth a higher dispersion of dissolved and particle-bound Pb and Znat that time as a consequence of removal and subsequently disposalof the former waste rock dump into the A-fiord as well as disposalof mine tailings with a higher soluble Pb content during the lastmonths of mine operation.
From 1991 to 2009, resident blue mussels at contaminated sitesin Maarmorilik showed decreasing trends in Pb and Zn concen-trations (Table 3). Decreases in Pb and Zn in resident blue musselswere observed at all blue mussel sites (a total of 19) included in theregular environmental program in Maarmorilik (Schiedek et al.,2009). For transplanted blue mussels, decreases in Pb contents(and Zn contents at Site T12SV nearest to the mine) were alsoobserved in the period 1991-2009. At the more distant sites, the Zncontamination in blue mussels was very low or negligible duringthe period when transplantation of blue mussels was conducted.Time trends for resident blue mussels after 1990 were significant atall sites at a 5% significance level and trends were predominatelylog-linear (exponential decreasing) (Table 3). The decrease in Pband Zn in resident and transplanted blue mussels is considered theresult of a decrease in dissolution of Pb and Zn from tailings andwaste rock in the A-fiord (now buried under a layer of naturalsediment) as well as a decrease in outwash of dissolved andparticle-bound Pb and Zn from the remaining land-based wasterock dumps nearMaarmorilik. For resident mussels, the decrease inPb and Zn was also influenced by disappearance of old mussels inthe size group preferred and the emergence of new generations ofmussels that had been exposed to lower concentrations.
In 2009, Pb contents in transplanted blue mussels within an areaof 12 km from Maarmorilik were still elevated by factors of 2e25after one year of transplantation (Fig. 8). In contrast, Zn levels wereless than a factor of 2 higher than the initial content for all sites(Fig. 8). The limited contamination by Zn in blue mussels in Maar-morilik is in contrast to seaweeds, which contain elevated concen-trations of Zn in a larger area around Maarmorilik. The observeddifferences between Zn in blue mussels and seaweed may beexplained by differences in sources ofmetal uptake between the two
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Site T12SV (<1 km from mine)
Site T12SV (<1 km from mine)
Site T17A (2 km from mine)
Site T17A (2 km from mine)
Site T36 (12 km from mine)
Site T36 (12 km from mine)
Fig. 7. Pb and Zn contents in resident and transplanted blue mussels at three sites in Maarmorilik with increasing distances to the mine sampled from 1985 to 2009. Transplantedmussels 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 contentshave been normalized to the contents in a 6 cm blue mussel (see Data analyses section in text). Each point represents the concentration in one subsample taken from a homog-enized bulksample containing approximately 20 indivial blue mussels.
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341 339
species i.e. metal uptake by bluemussels may bemore controlled byparticle-bound metals than dissolved metals. An alternative expla-nation is that blue mussels may have a greater capacity to regulatethe internal accumulation of Zn compared to seaweed, Zn being anessential nutritional element to biota in contrast to Pb.
4.6. Transplanted versus resident blue mussels
Transplanted blue mussels were used for monitoring recentmetal contamination at the coastal sites in addition to sampling ofresident blue mussels. A previous study from Maarmorilik showedthat once metals such as Pb and Zn were taken up in blue mussels,these were released at a very slow rate if the environmentalconditions improved (Riget et al., 1997a). Furthermore, blue musselsgrow relatively slowly in arctic waters and a blue mussel with thepreferred shell length of 6 cm is likely to be approximately 12 yearsold (Theisen, 1973). Consequently, resident blue mussels did notseem adequate for monitoring annual variations in seawatercontamination 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. Thedifferences between the Pb and Zn contents in transplanted versusresident bluemussels at typical sites can be seen in Fig. 7. The Pb andZn contents were consistently lower in transplanted mussels afterone year of transplantation compared to resident mussels. Thelargest differences between transplanted and resident mussels wereobserved just after the mining period when the change in contam-ination rate was highest and the concentrations in transplanted andresident blue mussels approached each other toward the end of themonitoring period. Also, it is notable that the Pb content in trans-planted mussels at Site T17A (Fig. 7) in 1985 were about 4 timeshigher than in 1992 while the resident mussels contained approxi-mately the same amount. The differences between Pb and Zn intransplanted versus resident blue mussels after 1990 show that Pband Zn are released from the resident mussels at a very slow rate. Italso shows that transplantation of blue mussels is an adequate wayto monitor short-term (annual) changes in seawater metalcontamination in contrast to resident mussels. This is at least thecase for monitoring metals such as Pb and Zn in arctic blue musselsin a situation where the contamination is decreasing.
0
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T5 (3 km)
T12SV(< 1 km)
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L, 2008 (35 km)
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Site
l e s s u m
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6 r e p g µ
Pb Zn
Fig. 8. Pb and Zn contents in blue mussels transplanted from Site L in 2008 andcollected at five sites near Maarmorilik in 2009. The distances from the mine areshown in parentheses. The metal contents have been normalized to the contents ina 6 cm blue mussel (see Data analyses section in text). Each bar represents theconcentration in one subsample taken from a homogenized bulksample containingtypically 20 individual blue mussels.
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341340
4.7. Contamination sources, processes and “lessons learned” inMaarmorilik
During themining period from 1973 to 1990, the major inputs ofmine-related metals to the marine environment in Maarmorilikcame from disposal of mine tailings and later waste rock into thesmall partly-enclosed A-fiord (Asmund, 1992b). The A-fiord isseparated from the outer Q-fiord by a shallow submarine sill and analmost stagnant water mass is created below the sill duringsummer. This presumably makes the A-fiord an ideal trap for sus-pended particles such as mine tailings, which at the time whenmining was commenced was thought to contain insoluble metalcompounds. However, as studies later revealed, the tailings con-tained heavy metals including Pb, Zn and Cd bound in chemicalforms that were indeed soluble in the ambient seawater. The mainminerals of the tailings were calcite, dolomite and pyrite but it wasindicated that most of the Zn and Cd were contained in the mineralsphalerite (ZnS). Pb was mainly contained in minerals other thansulphides, probably carbonates or sulphates (Loring and Asmund,1989; Asmund, 1992b). Studies showed that metals dissolvedduring the dilution of the tailings with the seawater and thesimultaneous decrease of pH from approximately 10 to 8 (Asmund,1992b). Seasonal mixing of the water column in the A-fiord waslater found to take place during winter as a result of freezing of thesurface water and an inflow of water between the A-fiord andQ-fiord (Møller, 1984). The complete vertical mixing of water andincreased water exchange between the two fiords increased thedispersion of dissolved metals as well as fine suspended tailingsparticles from the A-fiord to the Q-fiord. Also, the mixing of waterled to increased oxidation- and dissolution-potentials of the tail-ings particles in suspension as well as the tailings and waste rocksituated on the bottom of the A-fiord (Elberling et al., 2002). Duringthe mining period, the main process responsible for the inputof dissolved metals to the marine environment in Maarmorilikwas considered the dissolution of metals from mine tailings. Thedissolution process occurred mainly after discharge but beforesettling of the tailings close to the outlet point (Asmund, 1992b).This was shown by an abrupt decrease in Pb concentrations in thebottomwater of the A-fiord after the mine closed down in 1990 andthe last produced tailings had settled. While the main input ofdissolved metals came from the tailings, the waste rock dumpswere significant sources of contamination of the intertidal zones(Asmund, 1992b).
After 1990, Pb concentrations decreased abruptly in the bottomwater of the A-fiord while Zn concentrations did not decrease to thesame extent. During the last measurements in 2008, Zn levels werestill 50e100 times higher than in uncontaminated seawater. Thisindicates that considerable amounts of Znwere and still are releasedfrom the buried mine waste in the A-fiord while the dissolution ofPb are probably negligible. In contrast, the abrupt decrease in Pbconcentrations in water within the A-fiord was not reflected ina similar abrupt decrease in Pb contamination in seaweed and bluemussels. This indicates that themain sources of Pb contamination ofthemarine environment inMaarmorilik have likely shifted from thetailings disposed into the A-fiord to outwash of dissolved andparticle-bound Pb from the land-based waste rock dumps as well asfrom land-deposited dust that was spread during themining period.Given the time trends for Pb and Zn in seaweed and blue mussels, itseems likely that elevated levels of Pb and Zn in seaweed and bluemussels will still be measured for at least 50 more years nearMaarmorilik, even if no new mining activity is commenced.
In the Maarmorilik case, poor environmental studies and deci-sion-making led to significant metal contamination of the marineenvironment near Maarmorilik. Several lessons can be learnedfrom that, including those listed below. First of all, if mine waste isgoing to be disposed into the sea, very detailed studies of how thewaste will react in that environment has to be made prior todisposal because once it is disposed it is nearly impossible toretrieve. Studies prior to disposal must include representativewaste material (tailings/waste rock) obtained from a pilot projectprior to the mine operation subjected to the conditions in which itis going to be disposed (using representative seawater). More roughtests, such as leach tests using acetic acid will also provide valuableresults. The effects of different redox conditions on the chemicalbehavior of the mine waste have to be included in the tests as wellas a thorough understanding of the hydrological conditions all yearround at the disposal-site as these may effect the mobilization anddispersion of contaminants. Bioaccumulation studies are usefultools to reveal how different organisms react to the given exposure.During mining activities, attention must be given to minimize thegeneration and dispersal of dust as dust particles may remain in theterrestrial environment for several decades after the initial depo-sition. Finally, chemically reactive waste rock must be kept awayfrom the tidal zone and if at all possible transported to a confinedarea as part of the mining process and subsequently covered e.g.with chemically inert rock material such as the marble/dolomitefound in Maarmorilik. The remains of a former land-based wasterock dumpwere still a major contaminant source two decades aftermine closure, which shows that once large amounts of waste rockare dumped in an environmentally unsafe manner, it is extremelydifficult to remediate afterward probably due to finer grainedparticles remaining on site.
5. Conclusions
This study assessed the environmental state adjacent to theformer PbeZn mine in Maarmorilik, West Greenland two decadesafter mine closure. Despite the two decades; seawater, sedimentand biota, including seaweed and blue mussels were still contam-inated with Pb and Zn related to the past mining operation. Thisrevealed a slow recovery of this arctic fiord system to miningcontamination. Specifically, it was shown that 12 km from themine, seaweed and blue mussels still contained elevated concen-trations of Pb (and in seaweed also Zn). During the mining period,the primary source of contamination was considered the dissolu-tion of metals from mine tailings that were discharged into thesmall fiord Affarlikassaa near Maarmorilik. Later, following thesettling of tailings on the bottom of Affarlikassaa and the burial of
J. Søndergaard et al. / Marine Environmental Research 71 (2011) 331e341 341
these with natural sediment, the main sources of Pb contaminationof the marine environment in Maarmorilik likely changed. Twodecades after mine closure, remaining land-deposited waste rock,especially the residues of an old large-scale waste rock dump andoutwash of Pb-contaminated dust generated during the miningperiod were considered the dominant sources of Pb contaminationin Maarmorilik. The contamination by Zn, however, still seemed tobe considerably influenced by a release of Zn from the wastematerials at the bottom of Affarlikassaa.
Appendix A. Supplementary material
Supplementary data associated with this article can be found, inthe online version, at doi:10.1016/j.marenvres.2011.03.001.
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