Modern and historical fluxes of halogenated organic contaminants to a lake in the Canadian arctic, as determined from annually laminated sediment cores

Science of the Total Environment 342 (2005) 223 – 243 www.elsevier.com/locate/scitotenv

Modern and historical fluxes of halogenated organic contaminants to a lake in the Canadian arctic, as determined from annually laminated sediment cores G.A. Sterna,T, E. Braekevelta, P.A. Helma, T.F. Bidlemanb, P.M. Outridgec, W.L. Lockharta, R. McNeeleyc, B. Rosenberga, M.G. Ikonomoud, P. Hamiltone, G.T. Tomya, P. Wilkinsona a

Department of Fisheries and Oceans, Freshwater Institute, 501 University Crescent, Winnipeg, Manitoba, Canada R3T 2N6 Centre for Atmospheric Research Experiments, Meteorological Service of Canada, 6248 Eighth Line, Egbert, ON, Canada L0L 1N0 c Geological Survey of Canada, 601 Booth Street, Ottawa, Ontario, Canada K1A 0E8 d Department of Fisheries and Oceans, Institute of Ocean Sciences, 9680 West Saanich Rd., Sidney, British Columbia, Canada V8L 4B2 e Canadian Museum of Nature, P.O. Box 2443 Station D, Ottawa, Canada K1A 6P4 b

Abstract Two annually laminated cores collected from Lake DV09 on Devon Island in May 1999 were dated using 210Pb and Cs, and analyzed for a variety of halogenated organic contaminants (HOCs), including polychlorinated biphenyls (PCBs), organochlorine pesticides, short-chain polychlorinated n-alkanes (sPCAs), polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs), and polybrominated diphenyl ethers (PBDEs). Dry weight HOC concentrations in Lake DV09 sediments were generally similar to other remote Arctic lakes. Maximum HOC fluxes often agreed well with production maxima, although many compound groups exhibited maxima at or near the sediment surface, much later than peak production. The lower than expected HOC concentrations in older sediment slices may be due to anaerobic degradation and possibly to dilution resulting from a temporary increase in sedimentation rate observed between the mid-1960s and 1970s. Indeed, temporal trends were more readily apparent for those compound classes when anaerobic metabolites were also analyzed, such as for DDT and toxaphene. However, it is postulated here for the first time that the maximum or increasing HOC surface fluxes observed for many of the major compound classes in DV09 sediments may be influenced by climate variation and the resulting increase in algal primary productivity which could drive an increasing rate of HOC scavenging from the water column. Both the fraction ( F TC) and enantiomer fraction (EF) of trans-chlordane (TC) decreased significantly between 1957 and 1997, suggesting that recent inputs to the lake are from weathered chlordane sources. PCDD/Fs showed a change in sources from pentachlorophenol (PeCP) in the 1950s and 1960s to combustion sources into 137

T Corresponding author. Tel.: +1 204 984 6761; fax: +1 204 984 2403. E-mail address: [email protected] (G.A. Stern). 0048-9697/$ - see front matter. Crown Copyright D 2005 Published by Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2004.12.046

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the 1990s. Improvements in combustion technology may be responsible for the reducing the proportion of TCDF relative to OCDD in the most recent slice. Crown Copyright D 2005 Published by Elsevier B.V. All rights reserved. Keywords: Arctic; Organohalogens; Climate warming; Sediment cores; Anaerobic degradation; Temporal trends

1. Introduction Halogenated organic contaminants (HOCs), including industrial products such as polychlorinated biphenyls (PCBs) and organochlorine pesticides like DDT and toxaphene, are frequently found in the Arctic, despite never having been used there. Their presence implies that they are sufficiently persistent and volatile to be transported long distances from sources. Lake sediments are a sink for these compounds, and sediment cores have been used to determine modern and historical inputs of contaminants to lakes all over the world, including the Arctic (Lockhart et al., 1993, 1998; Muir et al., 1995, 1996; Tomy et al., 1999; Rawn et al., 2001). Unfortunately, most sediments are mixed to some degree by processes such as wave action, lake turnover, ice scouring, and the activity of benthic animals, hindering the elucidation of temporal trends in contaminant transport and deposition. Sediments that are deposited in discrete layers that have not been disturbed are rare. Cores taken from lakes with such sediments are valuable because their historical record is intact. Recently, Gajewski et al. (1997) reported a lake with laminated sediments near the northern coast of Devon Island (Lake DV09). We also collected sediment cores from Lake DV09 and analyzed them for a variety of HOCs, including PCBs, DDT, hexachlorocyclohexanes (HCHs), chlordane, toxaphene, endosulfan, short-chain polychlorinated n-alkanes (sPCAs), polybrominated diphenyl ethers (PBDEs), and polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/F). The major objective of this study was to determine modern and historic contaminant fluxes and inventories of these HOCs in Lake DV09 sediments with a higher degree of temporal resolution than would be possible in mixed sediments. In addition, the enantiomer composition of transchlordane (TC) in Lake DV09 sediments was compared to Arctic air samples collected in the early to late 1990s. Enantiomers are useful as marker com-

pounds to follow transport processes: microbially processed pesticides often have nonracemic proportions of enantiomers and can thereby be distinguished from the freshly applied racemic compounds (Bidleman et al., 2002, 2003).

2. Materials and methods 2.1. Sampling Lake DV09 (758 34.420V N, 898 18.545V W, 35 m a.s.l.) is a small (3 ha), shallow (maximum depth 13.5 m) lake situated near the northern coast of Devon Island in Nunavut Territory (Fig. 1). The lake is surrounded by till veneer on three sides and contacts an alluvial terrace on one side, and is underlain by Middle Ordovician Bay Fiord formation which consists of dolomite and gypsum with limestone, shale, and siltstone (Gajewski et al., 1997). An expedition to Lake DV09 was made in May 1999, using a ski-equipped Twin Otter from the Polar Continental Shelf Project in Resolute. The depositional zone of Lake DV09 was determined from bathymetric maps and was sampled with a 10 cm diameter KB corer. Four cores approximately 4 m apart were collected. The core tubes were carefully lowered to the lake bottom to avoid disturbing surface flocculent layers. The core tops were all undisturbed and distinct laminations were observed down the lengths of each. Cores 1 and 2 were sliced into 0.5 (surface layers) and 1 cm (deeper layers) intervals on site, immediately placed into plastic (WhirlpakR) bags, weighed, sealed, and stored at ambient temperature (0–10 8C) after collection. Cores 3 and 4 were sealed in the core tube with plastic and styrofoam plugs and also stored at ambient temperature on site. All samples from the two sliced cores were brought to the Freshwater Institute for dating and contaminant analysis while the two intact cores were sent to the Geological Survey

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225

100 km

75° N

DEVON ISLAND

60°N 1 km

CANADA

100 m

75° 34.420' N 89° 18.545' W 35 m a.s.l.

LAKE DV09

200 m 100 m

200 m

Fig. 1. Map showing location of Lake DV09, Devon Island, Nunavut, Canada (Gajewski et al., 1997).

laboratories in Ottawa for geological and geochemical analyses and diatom enumeration (Outridge et al., in press). After transport to the laboratory, wet sediments were freeze-dried in their sample bags and stored at 5 8C in the dark until they were sub-sampled for analysis of 210Pb, 137Cs, metals, HOCs, total, and inorganic carbon. Organic carbon was determined by combustion of dried sediment in an oxygen–helium atmosphere at 950–975 8C and quantification of CO2 using a CE 240-XA Elemental Analyser (Exeter Analytical, North Chelmsford, MA) and was strongly associated with total diatom abundances (Outridge et al., in press). 2.2. HOC analysis Core 1 was analyzed for PCBs, OC pesticides, and sPCAs. Freeze-dried sediments were combined with

anhydrous sodium sulfate and extracted in an accelerated solvent extractor (ASE 200, Dionex Canada, Oakville, ON). PCB 30 and octachloronaphthalene recovery standards were added before extraction. Sulfur was removed by treatment of the extracts with activated copper powder (Muir et al., 1996). Extracts were reduced in volume and fractionated on 1.2% deactivated Florisil (Norstrom et al., 1988). Fractions were then analyzed by high-resolution gas chromatography with electron capture detection (HRGCECD), using a 60 m DB-5 capillary column (0.25 mm ID, 0.25 Am film thickness, J&W Scientific). Chemicals were quantified by comparing chemical peak areas (identified by retention time) to those of commercially available standards of known concentration. Trans-chlordane (+) and (
) enantiomers were separated by chiral stationary-phase capillary gas chromatography (GC) coupled with electron

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capture negative ion low resolution mass spectrometry (ECNI-LRMS). An enantiomer fraction (EF) was calculated as (+)/[(+)+(
)]; a racemate has equal amounts of each enantiomer and an EF of 0.500 (Bidleman et al., 2002). For sPCA and toxaphene analysis, all fractions were combined. A known amount of [13C8]mirex was added as an internal standard prior to analysis to correct for variations in MS performance between injections. Samples were then analyzed by GC combined with electron capture negative ion (ECNI) high resolution mass spectrometry (HRMS) (Tomy et al., 1997; Braekevelt et al., 2001). Core 2 was analyzed for PCDD/Fs and PBDEs. Sediment samples, approximately 5–10 g wet weight, were combined with Na2SO4, spiked with a mixture of 13C-labelled PCDD/Fs and PBDEs (Cambridge Isotope Laboratories, Andover, MA), and Soxhlet extracted for 16 h with 350 mL of toluene/acetone (80:20). Extracts were washed sequentially with 40 mL KOH (1 M), 80 mL HPLC grade water, and 10 mL of H2SO4, then rotary evaporated and reconstituted in 10 mL of DCM/hexane (1:1). Sample cleanup steps consisted of a silica gel column (with layers of basic, neutral, acidic, neutral silica), a column filled with copper filings and Na2SO4 (to remove sulfur and residual water), and an activated neutral alumina column capped with anhydrous Na2SO4. Extracts were then fractionated on a carbon-fiber column connected to an automated high performance liquid chromatography (HPLC) system. Four fractions were collected: the first three fractions contained PCBs and PBDEs. Fraction IV, containing PCDD/Fs, was concentrated to 10 AL, spiked with 13C-labeled PCDD/F performance standards and analyzed by HRGC combined with electron impact (EI)-HRMS. All fractions were then combined, reduced in volume to 10 AL, spiked with 13C-labeled PBDEs method performance standards, and analyzed by HRGC-EIHRMS for PBDEs (Ikonomou et al., 2001). 2.3. Quality assurance/quality control Precautions taken during analysis to reduce background contamination include the use of glassdistilled solvents and baking of glassware, Na2SO4, and Florisil. Procedural blanks, which were taken through all phases of extraction and cleanup, were

included with each batch of samples. In addition, several pre-1900 slices were analyzed from each core. Concentrations were blank-corrected by subtracting the mean values of individual compounds measured in the pre-1900 slices from each of the sample slices. Quantitative results are regularly compared with other laboratories through participation in interlaboratory exercises for the analysis of chlorinated contaminants in a variety of environmental matrices and through the use of standard reference materials. 2.4. Sediment core dating, focusing factors, sedimentation rates, and flux calculations Core slices were dated by measuring 210Pb and Cs activity as a function of depth. Sedimentation rates were calculated using linear and constant rate of supply (CRS) models (Robbins, 1978). Precision of the CRS dates and sedimentation rates were calculated using the method of Wilkinson and Simpson (2003). 137 Cs activity (from the atmospheric testing of nuclear weapons, which began in 1954 and peaked in 1963) was also measured to confirm 210Pb dates. Sediment focusing factors were calculated as the ratio of the sediment 210Pb flux to the atmospheric 210Pb flux, which was estimated from a soil profile taken from the north shore of Banks Island in 1998 (74 8N). Fluxes were corrected for differences in both organic carbon and sedimentation rates among slices, which gives a more accurate estimation of HOC inputs to the lake by reducing variability due to changes in watershed processes: 137

Flux ¼

Cdw Sdw foc F

where S dw is the sedimentation rate in each slice (g m
2 y
1), C dw is the (blank-corrected) dry-weight HOC concentration in each slice (ng g
1), f oc is the fraction of organic carbon in each slice, and F is the focusing factor of the core. Dry-weight contaminant fluxes and HOC inventories were also determined to allow comparison with literature values for other lakes: X Cdw Sdw I Inventory ¼ F where I is the slice interval in years (Muir et al., 1996).

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227

ing of sediment to both collection sites. Focusing factors for the two cores were 2.40 and 2.11, respectively. Dating of the sediment slices with 210Pb coincided with counts of laminations. In addition, the profile of 137 Cs activity agreed well with 210Pb dates (Fig. 2). The shoulder in the 137Cs profile (dated 1988-core 1) probably represents a small input from the Chernobyl reactor accident in the Ukraine in 1986, and has been observed elsewhere (Lockhart et al., 1993). There was good agreement between the two cores dated in this study and those studied by Gajewski et al. (1997). Sedimentation rates, calculated using a linear model, were 300 and 290 g m
2 y
1 for the two cores, which compare very well with the sedimentation rate of 295 g m
2 y
1 reported by Gajewski et al. (1997). There was also good agreement between the linear and constant flux models (267 and 278 g m
2 y
1 for cores 1 and 2, respectively) used to date the core and determine sedimentation rates, and good agreement in sedimentation rates and percent organic

2.5. Diatom enumeration Diatom concentrations were determined at the Canadian Museum of Nature using methods described by Gajewski et al. (1997). Results from this study are outlined by Outridge et al. (in press).

3. Results and discussion 3.1. Core dating, sedimentation rates, and focusing factors All cores were annually laminated. The varves were clearly visible in X-ray images (Gajewski et al., 1997; Outridge et al., in press). Laminated couplets consisted of a light inorganic layer, representing clastic deposition from allochthonous sources, alternating with a darker biogenic layer deposited in the summer. The top 9 cm represented approximately 100 years of sedimentation. There was some focus-

Excess 210Pb (Bq g-1) 0.001 0

0.01

0.1

1

0.01

0.1

Cumulative dry weight (g cm-2)

1988

1

1 0

1997

1997

1986

1971

1

1970 1964

1963

1940

1938

2

2 137

Cs 210 Pb

1908

1910

3

3

Core 1

4

m-2

Core 2

y-1

Avg. sed. rate = 278 g m-2 y-1 210 Pb flux = 57 Bq m-2 y-1

Avg. sed. rate = 267 g 210 Pb flux = 65 Bq m-2 y-1

5 0.00

0.05

0.10

0.15 137

4

0.06

0.12

5 0.18

Cs (Bq g-1)

Fig. 2. 210Pb and 137Cs activities, the estimated dates of slice mid-points in DV09 cores 1 and 2 and their corresponding sedimentation rates calculated using the constant flux model (Wilkinson and Simpson, 2003).

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carbon between the cores (Fig. 3). A temporary increase in sedimentation rate and a corresponding decline in organic carbon was observe in the early 1970s for cores 1 and 2 of this study as well as the DV09 core collected by Gajewski et al. (1997) in 1994. The sedimentation rate measured in Lake DV09 is similar to those of other nearby lakes (Muir et al., 1995; Lockhartm et al., 1998).

3.2. HOC trends In general, dry weight HOC concentrations in Lake DV09 sediments (Table 1) were similar to other remote Arctic lakes (Muir et al., 1995, 1996; Gubala et al., 1995; Malmquist et al., 2003). Lake sediments from Schrader Lake in Alaska were very low in HOCs compared to Lake DV09 (Gubala et al., 1995). This

A 1 2 3 4

Slice

5 6 7 8 Core 1 Core 2

9 10 11 12 150

200

250

300

350

400

450 -2

500

550

-1

Sedimentation rate (g m y )

B

1 2 3 Core 1 Core 2

4

Slice

5 6 7 8 9 10 11 12 0.8

1.0

1.2

1.4

1.6

1.8

2.0

2.2

2.4

2.6

% organic carbon Fig. 3. (A) Sedimentation rates and (B) % organic carbon in DV09 cores 1 and 2.

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Table 1 Dry weight concentrations (maximum and surface), focus corrected fluxes (maximum and surface), and inventories of HOCs in Lake DV09 sediment cores Compound group

Concentration (ng g
1, dry weight)

Flux (ng m
2 y
1, dry weight)

Maximum (year)

Surface

Maximum (year)

Surface

Inventory (Ag m
2)

ADDT a p, pV-DDT p, pV-DDE p, pV-DDD AHCHb a-HCH h-HCH g-HCH ACHLc cis-chlordane trans-chlordane cis-nonachlor trans-nonachlor APCBd AMono- and Di-CB ATri-CB ATetra-CB APenta-CB AHexa-CB AHepta-CB AOcta-CB ACHB AsPCAe Endosulfan Dieldrin APCDFf APCDDf PCDD/F TEQf APBDE

0.20 0.01 0.14 0.02 0.33 0.22 0.06 0.17 0.20 0.05 0.03 0.03 0.06 2.68 0.09 0.98 0.73 0.56 0.48 0.19 0.05 1.47 17.6 0.04 0.43 2.11 2.35 0.07 0.25

0.05 0.01 0.04 0.01 0.22 0.12 0.02 0.08 0.17 0.04 0.01 0.03 0.05 2.56 0.03 0.67 0.73 0.56 0.36 0.19 0.01 0.72 17.6 0.04 0.28 0.43 0.86 0.005 0.17

15.5 1.2 10.8 3.3 37.9 25.7 4.7 12.9 24.2 5.9 3.6 4.0 7.2 370 10.7 115 105 80.9 52.4 27.2 4.8 166 2.5 6.2 46.1 279 248 8.7 28.5

7.1 1.2 5.1 0.8 31.3 17.9 2.5 10.9 24.2 5.6 2.0 4.0 7.2 370 3.9 97 105 81 52 27 2.0 103 2.5 6.2 40.5 72.2 144 0.9 28.5

0.60 0.05 0.42 0.09 1.2 0.61 0.11 0.47 0.66 0.15 0.12 0.07 0.17 11.8 0.30 4.3 2.7 2.2 1.5 0.56 0.20 3.9 49 0.04 1.9 8.3 9.0 0.14 1.1

(1957) (1988) (1957) (1964) (1957) (1993) (1957) (1957) (1988) (1988) (1957) (1997) (1988) (1993) (1993) (1993) (1997) (1997) (1952) (1997) (1952) (1980) (1997) (1997) (1964) (1978) (1963) (1978) (1951)

(1957) (1997) (1971) (1971) (1993) (1993) (1993) (1957) (1997) (1993) (1971) (1997) (1997) (1997) (1993) (1993) (1997) (1997) (1997) (1997) (1971) (1980) (1997) (1997) (1971) (1970) (1963) (1978) (1997)

a

Sum of p, pV-DDT, p, pV-DDE, p, pV-DDD, o, pV-DDT, o, pV-DDE, and o, pV-DDD. Sum of the a-, h-, and g-HCH isomers. c Sum of all chlordane related compounds, including heptachlor epoxide. d Sum of CB1, 3, 4/10, 7, 6, 8/5, 19, 18, 17, 24/27, 16/32, 26, 25, 31, 28, 33, 22, 45, 46, 52, 49, 47, 48, 44, 42, 41/71, 64, 40, 74, 70/76, 66, 95, 56/60, 91, 84/89, 101, 99, 83, 97, 87, 85, 136, 110, 82, 151, 144/135, 149, 118, 134, 114, 131, 146, 153, 132, 105, 141, 130/176, 179, 137, 138, 158, 178/129, 175, 187, 183, 128, 185, 174, 177, 171, 156, 201/ 157, 172/197, 180, 193, 191, 200, 170, 190, 198, 199, 196/203, 189, 208, 195, 207, 194, 205, 206, 209. e Flux and inventory units are in Ag m
2 y
1 and mg m
2, respectively. f Concentration, flux, and inventory units are in pg g
1, pg m
2 y
1, and ng m
2, respectively. b

may be attributed to its low sedimentation rate (55 g m
2 y
1), or to the prevailing weather patterns, which isolate western sinks from eastern sources (Commoner et al., 2000). Many of the HOCs examined, including PCBs, PBDEs, HCHs, sPCAs, and endosulfan, exhibited maxima at or near the sediment surface, despite having considerably earlier production maxima. Anaerobic degradation was found to be an important

removal process for some HOCs in Arctic lake sediments (Muir et al., 1995), and may explain the lower concentrations observed for a number of compound classes in older sediment slices. However, compound classes such as DDT and toxaphene degrade to metabolites that were also analyzed (e.g., DDD and DDE); as a result, the totals (parent + metabolites) did not change with time and temporal trends were more readily apparent.

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Watershed processes such as variations in ice cover, productivity, and precipitation which can be significantly impacted by climate variation may also influence contaminant inputs to Arctic lake systems (Helm et al., 2002; Macdonald et al., 2004). Indeed, warming has significantly increased the rate of spring thaw (Magnuson et al., 2000) and phytoplankton productivity (Douglas et al., 1994; Wolfe, 2000). The observed occurrence of HOC flux maxima and secondary maxima in the surface slices of the DV09 cores may in part be influenced by the significant limnological changes which the lake has undergone over the past century and which appear to be climaterelated (Gajewski et al., 1997; Outridge et al., in press). This is illustrated for mercury (Outridge et al., in press) and in this paper for HOCs by including diatom concentration trends as an indicator of lake productivity changes. 3.2.1. Polychlorinated biphenyls (PCBs) Recently, Breivik et al. (2002) published the results of a study designed to quantitatively estimate the historical production of selected PCB homologue groups and individual congeners. Total global PCB production was estimated to be 1324 kt between 1930 and 1993: the tri-CBs accounted for 27.8% of total production, followed by the tetra-, penta-, and hexa-CBs at 25.2%, 19.4%, and 11.3%, respectively. As a result of an increased awareness of the environmental effects of the more highly chlorinated PCB homologues and congeners, the production of the heavier technical mixtures and homologue groups decreased in the last few decades of production. Worldwide PCB production is thought to have ended when Russia ceased production of Sovol in 1993. Total PCB flux peaked in the top two slices (1993– 1997) with smaller maxima occurring in the core slices dated to 1971, during peak PCB production, and prior to increased production in 1952 (Fig. 4A). Anaerobic degradation and possibly dilution due to a temporary increase in sedimentation rate observed around 1971 (Fig. 2) can both contribute to the lower PCB flux observed during peak PCB production relative to the surface sediment. However, the overall pattern of PCB flux is very similar to that shown by total diatom concentrations which peaked first between the early 1940s and the mid-1950s and then

increased exponentially from about 107 to 109 cells/g DW after about the early 1980s. Similar concentrations to those observed in Lake DV09 have been reported for a number of other Arctic lakes (Gubala et al., 1995; Muir et al., 1996; Vartiainen et al., 1997; Malmquist et al., 2003). The focus-corrected APCB inventory for Lake DV09 (9.8 Ag m
2, Table 1) compares well with those of Amituk (8.7 Ag m
2) and Hazen (9.5 Ag m
2) Lakes in the Canadian Arctic (Muir et al., 1996), although Lake DV09 has approximately 10 more years of PCB accumulation. Maximum APCB concentrations in Amituk and Hazen occurred in the surface sediment dated to 1986, and decreased steadily with depth over the 20th Century (Muir et al., 1996). An increase in diatom concentrations, similar to that observed in DV09, was also observed in an Amituk Lake sediment core (Outridge et al., in press). The PCB homologue profiles in DV09 were dominated by the tri- and tetra-CBs, as might be expected based on the global production estimates of PCB homologues described above. However, the relative abundance of the tri-CBs declined from the onset of PCB production to 1997, whereas the proportion of the tetra- and penta-CBs increased (Fig. 4B). No significant changes were observed for the mono-, di-, and hexa- to octa-CBs. A decrease in the proportion of lower chlorinated (tri- and tetra-) PCBs was also observed in arctic air over the period 1993–1997, but there was no apparent trend for the heavier PCB congeners (Hung et al., 2001). 3.2.2. Toxaphene (RCHB) Toxaphene (chlorobornanes, ACHB) was introduced in the U.S.A. in 1945 by Hercules as a new insecticide to control a variety of insect pests. While over two-thirds of the total production was used for insect control on cotton in the southern U.S. (Li, 2001), it was also used on vegetables, small grains, soybeans, for control of external insects on livestock as well as a piscicide to rid lakes of undesirable fish species (Miskimmin et al., 1995). Toxaphene was deregistered in the U.S. for most applications in 1982 and was withdrawn from all applications in 1986 (USEPA, 1982). However, toxaphene continues to be produced and used in many other countries (Voldner and Li, 1995), and is ubiquitous throughout the arctic ecosystem (Barrie et al., 1992).

G.A. Stern et al. / Science of the Total Environment 342 (2005) 223–243

A

231

Diatoms (cells/g DW) 1e+5 2000

1e+6

1e+7

1e+8

1e+9

1e+10

1980

Year

1960

1940 ΣPCB Flux Diatoms Production

1920

1900 0

5

10

15

20

25

ΣPCB Flux (µg m-2y-1) 0

50

100

150

200

250

300

Production (kt)

B 60

% ΣPCB

50

Tri-CBs Tetra-CBs Penta-CBs

Tri-CBs (R2 = 0.42, p < 0.02)

40

30

Tetra-CBs (R2 = 0.37, p < 0.03) 20 Penta-CBs (R2 = 0.43, p < 0.02) 10 1920

1930

1940

1950

1960

1970

1980

1990

2000

Year Fig. 4. (A) Lake DV09 APCB organic carbon and focus corrected flux, estimated global production (Breivik et al., 2002) and the historical diatom concentration trend (Outridge et al., in press); (B) percentages of selected major PCB homologue groups.

Toxaphene concentrations, focus corrected fluxes, and inventories in Lake DV09 sediments (Table 1) were about an order-of-magnitude lower than those in sediment cores from the Great Lakes (Pearson et

al., 1997; Schneider et al., 2001). Maximum flux occurs in the slice dated to 1980 (Fig. 5B), which is approximately 6 years after peak toxaphene production (Rapaport and Eisenreich, 1988; Voldner and Li,

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A

Diatoms (cells/g DW) 1e+5 2000

1e+6

1e+7

1e+8

1e+9

1e+10

1980

Year

1960

1940 ΣCHB Flux Diatoms Input function

1920

1900 0

3

6

9

12

15

ΣCHB Flux (µg m-2y-1)

B

70 60

% ΣCHB

50 hexa hepta octa nona

40 30 20 10 0 1920

1940

1960

1980

2000

Year Fig. 5. (A) Lake DV09 toxaphene (ACHB) organic carbon and focus corrected flux (sum of hexa- to nonachlorobornanes) and the historical diatom concentration trend (Outridge et al., in press). The curve corresponds to the atmospheric input function of toxaphene derived by Rapaport and Eisenreich (1988); (B) relative proportions of hexa- to nonachlorinated homologue groups.

1995). However, as stated above, toxaphene usage was not completely banned in the U.S. until the early mid-1980s and continued to be produce and used in other countries into the 1990s. An increase in ACHB flux was observed in the top two slices of the core, which like PCBs may be attributed to increasing diatom concentrations. However, unlike the PCBs,

the maximum flux was still observed near peak production and before the observed increase in diatom concentrations in the late 1980s. This as stated earlier, and as is discussed below, may be attributed to the fact that both parent and metabolite toxaphene congeners were analyzed in each core slice.

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The heptachlorobornane congener B7-1001 is predominant throughout the core, comprising 45 to 59% of ACHB. The proportion of hexachlorobornanes increased with sediment age, whereas the other examined homologue groups decreased (Fig. 5B). This is due to anaerobic degradation of the more heavily chlorinated congeners to the dead-end metabolite B6-923, a hexachlorobornane (Stern et al., 1996; Ruppe et al., 2003). The doubling time for the hexachlorobornanes was calculated to be 41 years (R 2=0.91, pb0.001) and the degradation half-lives for the hepta-, octa-, and nonachlorobornanes were 190 years (R 2 =0.35, p=0.042), 76 years (R 2 =0.91, pb0.001), and 24 years (R 2=0.59, p=0.13), respectively. The longer half-lives of the lower chlorinated homologue groups are probably a result of stepwise degradation of the higher chlorinated congeners, and the heptachlorobornane half-life is particularly high because the congener B7-1001 is also a metabolite (Stern et al., 1996; Ruppe et al., 2003). 3.2.3. DDT and metabolites Maximum flux of ADDT occurred in the slice dated to 1971, approximately 8 years after peak production (Fig. 6A). The ADDT concentration in the surface sediment is still well above the 1950 levels, which again may be attributed to increased diatom concentrations and also to the fact that this pesticide is still being used both legally and illegally in countries around the world (Voldner and Li, 1995). Like toxaphene, both parent and degradation products were analyzed in each of the core slices. As is observed in the arctic atmosphere (Hung et al., 2002a), the major contributor to ADDT in the DV09 sediment was p,pV-DDE. The relative abundance of p,pV-DDE was highest near the sediment surface, whereas p,pV-DDD showed the opposite trend (Fig. 6B). This suggests that p,pV-DDE is the primary DDT component deposited to the lake surface, but some is degraded to the anaerobic metabolite p,pVDDD as it is buried. p, pV-DDT also appears to undergo anaerobic degradation: the half-life of p,pVDDT (using the ratio p, pV-DDT/( p,pV-DDD+ p,pVDDT)) in Lake DV09 sediments was calculated to be 22 years (R 2=0.88, p=0.005). Rawn et al. (2001) reported similar half-lives of 32 and 21 years for two Yukon Lakes, while Muir et al. (1995) reported a much longer half-life of 74 years for the anaerobic

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conversion of p,pV-DDT to p,pV-DDD in sediments from Amituk Lake on nearby Cornwallis Island. 3.2.4. Endosulfan Endosulfan is currently used in the United States, Canada, and other countries for insect control on high value crops. Annual usage of endosulfan in the U.S. in 1997 was approximately 725,000 kg (NCFAP, 2000). Technical endosulfan contains two isomers, endosulfan I, and endosulfan II, in approximately a 7:3 ratio along with impurities and degradation products (Rice et al., 1997). As was observed in arctic air (Hung et al., 2002a), only endosulfan I was present in the sediment from Lake DV09. The concentration of endosulfan I was highest at the sediment surface, and rapidly decreased to below detection limits in core slices dated prior to 1988. The rapid decline in endosulfan I with sediment age may be due to abiotic and/or biotic degradation, as it has been found to readily undergo both chemical and biological hydrolysis (Cotham and Bidleman, 1989; Guerin, 1999; Walse et al., 2002). In nearby Amituk Lake, overwinter water concentrations of endosulfan I decreased to below detection, and model predictions attributed these observations to hydrolysis (Helm et al., 2002; Diamond et al., 2005). A surface maximum in DV09 is also consistent with the recent increases in lake productivity, an increase in endosulfan I observed in Alert air over the period from 1993 to 1997 (Hung et al., 2002a), and with the 3.2-fold increase of endosulfan sulfate, the aerobic degradation product of endosulfan I, measured in Cumberland Sound beluga blubber over the 20 year time period from 1982 to 2002 (Stern and Ikonomou, 2003). 3.2.5. Chlordane Technical chlordane was used agriculturally to control cutworms in corn and in urban areas for termite control. It has been estimated that over 70,000 tons of technical chlordane has been produced since 1946 (Dearth and Hites, 1991). Although chlordane use has been banned in the U.S., Canada, and most European countries since the 1980s, it continued to be produced and exported until 1997 (PANNA, 1997). Total chlordane (ACHL) flux peaked in the core slice dated to 1971 and again in the surface sediment (Fig. 7A). Cis- and trans-chlordane (CC, TC) and cisand trans-nonachlor (CN, TN) were the predominant

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A

Diatoms (cells/g DW) 1e+5 2000

1e+6

1e+7

1e+8

1e+9

1e+10

1980

Year

1960

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ΣDDT Flux Diatoms Production

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1900 0.0

0.5

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ΣDDT Flux (µg m-2y-1) 0

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100

80

100

U.S. production (kt)

B

2000

Year

1990

1980 % pp-DDE % pp-DDD 1970

1960 0

20

40

60

% ΣDDT Fig. 6. (A) Lake DV09 ADDT organic carbon and focus corrected flux, the historical diatom concentration trend (Outridge et al., in press) and estimated historical rates of U.S. DDT production (Rapaport et al., 1985); (B) Lake DV09 percentages of p,pV-DDE and p,pV-DDD.

chlordane compounds present in the sediments, together accounting for ~75% of ACHL. The fraction of TC (FTC=TC/(TC+CC)) decreased significantly between 1957 and 1997 (FTC=m*year+b, m=
0.0075, b=15.3, R 2=0.63, pb0.04). A similar decrease was

seen in arctic air, where Bidleman et al. (2002) reported FTC slopes of
0.0079 and
0.0089 and intercepts of 16.1 and 18.1 for the period 1984–1998. FTC in the DV09 core slice dated to 1957 was 0.58, which is very similar to the FTC of technical chlordane

G.A. Stern et al. / Science of the Total Environment 342 (2005) 223–243

A

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Diatoms (cells/g DW) 1e+5 2000

1e+6

1e+7

1e+8

1e+9

1e+10

1980

Year

1960

1940

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ΣCHL Flux Trans-CHL Flux Diatoms

1900 0.0

0.4

0.8

1.2

1.6

Flux (µg m-2y-1)

B

0.51 0.50

0.5 = Racemic EF Sweden, Iceland, Slovakia 1971-73

EF

0.49 0.48

Finland 1998

0.47 0.46 0.45 0.44 1930

Russia & Canada 1994

EF = -0.0046 (year) + 1.4 R2= 0.81, p < 0.001)

1940

1950

1960

1970

1980

1990

Sweden 1998

2000

2010

Year Fig. 7. (A) Lake DV09 total chlordane (ACHL) organic carbon and focus corrected flux and the historical diatom concentration trend (Outridge et al., in press); (B) trans-chlordane enantiomer fraction (EF).

(0.61, Bidleman et al., 2002). In contrast, FTC in the most recent DV09 core slices were 0.27–0.33. These lower FTC values are similar to recent FTC values in arctic air and are characteristic of recycled inputs, suggesting that early inputs to the lake were from atmospheric transport and deposition of freshly applied chlordane, whereas more recent inputs are from weathered chlordane sources. This result is also

reflected in the fact that trans-chlordane did not show a corresponding increase with diatom concentrations as did the other major chlordane congeners. Enantiomer fractions (EFs) of TC were highest in sediments deposited in the 1940s and 1950s, and decrease near the surface: there is a strong negative correlation between TC EF and sediment age (Fig. 7B). More recent DV09 EF values compare well with

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those in arctic air from Rorvik, Sweden in 1998 and from Dunai, Russia and Alert, Canada in 1994, further suggesting that recent chlordane deposition to Lake DV09 is from a weathered source. However, older DV09 slices have lower EF values than the near racemic EFs in archived 1971–73 European air samples (Fig. 6B, Bidleman et al., 2003). This may be due to microbial degradation of the (+) TC enantiomer in the sediment as compared to archived air samples. EFs do not necessarily reflect the extent of total degradation, as considerable degradation may have occurred before deposition, and some HOCs can be readily degraded non-enantioselectively (Helm et al., 2000). 3.2.6. Hexachlorocyclohexanes (HCHs) Usage of technical HCH (60–70% a-HCH, 5–12% h-HCH, and 10–15% g-HCH) as a pesticide has dropped considerably in recent years as a result of restrictions in China in 1983 and India and the former Soviet Union in 1990–1992. Arctic air concentrations of a-HCH have reflected this decline in usage (Li et al., 1998), whereas g-HCH (lindane), which is still used in Europe, North America, and Asia, has declined only slightly in recent years (Hung et al., 2002a). The HCH flux indicates an influence of production in conjunction with other processes (Fig. 8A). Two maxima in the slices dated 1957 and 1971 correspond approximately with technical HCH production in the United States and China, respectively, while Europe was also shown to be a major contributor to the historical loading of a-HCH to the Arctic (Toose et al., 2004). However, there is also considerable HCH flux near the sediment surface. Anaerobic degradation may explain the lower HCH concentrations in older sediment slices while the increasing HCH surface flux may be attributed, at least in part, to the effects of climate warming and the resulting increase in primary productivity. a-/g-HCH ratios in the DV09 surface sediments (1.2–3.5, Fig. 8B) are similar to those reported in other arctic lake sediments (0.6–3.3, Muir et al., 1995) and arctic water (3.4F1.3, Harner et al., 1999; 3.4– 5.5, Law et al., 2001). However, in Lake DV09, the a-/g-HCH ratio decreases with sediment age. This is the opposite of the expected trend: a-/g-HCH ratios should be higher in the older sediments, which were

deposited at a time when technical HCH was more widely used, and decrease near the surface, when lindane (g-HCH) use became more widespread. It appears that a-HCH is preferentially degraded in Lake DV09 sediments. a-HCH was degraded more rapidly than g-HCH in Amituk Lake (Helm et al., 2000, 2002; Diamond et al., 2005) and the eastern Arctic Ocean (Harner et al., 1999). a-HCH hydrolyzes more rapidly (Ngabe et al., 1993; Harner et al., 1999) and degrades aerobically at a faster rate than g-HCH (Bachmann et al., 1988). These processes may contribute to the lower a-/g-HCH ratios observed in the older Lake DV09 sediments. However, under anaerobic conditions, g-HCH is less stable than a-HCH (Buser and Mu¨ller, 1995; Jagnow et al., 1977), and isomerization processes tend to favour the formation of a-HCH (Benezet and Matsumura, 1977; Vonk and Quirijns, 1979), which is contrary to the trend towards lower a-/g-HCH ratios in the sediments. Degradation of aHCH at a higher rate than g-HCH is the likely explanation but further study on degradation of HCHs is needed to determine the processes that result in the observed trends. 3.2.7. Short-chain polychlorinated n-alkanes (sPCAs) Also known industrially as chlorinated paraffins, polychlorinated n-alkanes (PCAs) are produced by direct free radical chlorination of n-alkane feedstocks with molecular chlorine. These reactions, which have low positional selectivity, yield extremely complex mixtures of optical isomers and congeners (Tomy et al., 1997; Tomy and Stern, 1999). Based on the principal n-alkane feedstocks, commercial PCA formulations fall into three categories: C10–C13 (short), C14–C17 (medium), and C20–C30 (long). PCAs are most commonly used as high temperature lubricants in metal working machinery and as flame-retardant plasticizers. Although the global production of sPCAs has been declining since the early 1980s, they have been measured in a wide range of environmental compartments including marine mammals, sediment, and air from the Canadian arctic (Muir et al., 2000a,b; Alaee et al., 2003; Stern and Evans, 2003; Stern, unpublished data). As a result, sPCAs have been placed on the Environmental Protection Agency’s Toxic Release Inventory, and have been classified as btoxicQ under the Canadian Environmental Protection Act.

G.A. Stern et al. / Science of the Total Environment 342 (2005) 223–243

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Diatoms (cells/g DW) 1e+5 2000

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1e+7

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Chinese production (kt) 0

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U.S. production (kt)

B

2000

1990

Year

1980

1970

1960

1950

1940 0

1

2

3

4

α–HCH/γ-HCH Fig. 8. (A) Lake DV09 AHCH organic carbon and focus corrected flux, the historical diatom concentration trend (Outridge et al., in press) and estimated U.S. (Rapaport and Eisenreich, 1988) and Chinese (Li et al., 1998) production; (B) Lake DV09 a-HCH/g-HCH ratios.

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Maximum sPCA flux occurs in the surface sediment, with a smaller peak in the core slice dated to 1957 (Fig. 9). Concentrations are similar to those in other arctic lakes (Tomy et al., 1999). Although the general pattern is consistent with historical production, maximum flux to DV09 occurred at least 10 years later than expected. Once again, the occurrence of near surface flux maxima and the core flux profile may be influenced by changes in the loading resulting from increases in primary production as well as degradation of sPCAs in the older slices. Shorter carbon chain length and lower chlorinated C10 and C11 formula groups become more predominant as you move down the core suggesting microbial degradation of the longer chain, more highly chlorinated compounds.

combustion sources is dominated by lower-chlorinated PCDFs, particularly the TCDFs. In contrast, the predominance of higher chlorinated PCDDs (particularly OCDD) is indicative of a pentachlorophenol (PeCP) source (Hagenmaier et al., 1994; Kjeller et al., 1996). PeCP was commonly used as a wood preservative from 1930 until the 1980s, peaking in the 1970s (WHO, 1987). OCDD is a major impurity in PeCP technical mixtures and can also be formed by photolytic degradation of PeCP (Crosby et al., 1981). APCDD/F levels in the DV09 core peaked in 1970, which was also observed in the Great Lakes (Pearson et al., 1995). The PCDFs started to increase in the 1950s and peaked in 1970, whereas the PCDDs began to increase about 10 years earlier and remained relatively constant from the 1960s to the most recent core slice (Fig. 10A). APCDD/F concentrations were unsurprisingly much lower than those in sediments from the Great Lakes (Pearson et al., 1995), but were also considerably lower than those in lake sediments from northern Finland (Vartiainen et al., 1997). The difference is probably due to the proximity of the Finnish lakes to sources in Europe (less than 2000

3.2.8. Polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/F) The combustion of residential waste in commercial incinerators is the largest single source of PCDD/F emissions in industrialized countries (Commoner et al., 2000). The PCDD/F homologue profile from these

Diatoms (cells/g DW) 1e+5 2000

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1e+8

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1e+10

1980

Year

1960

1940

ΣsPCA Flux Diatoms Production

1920

1900 0

20

40

60

80

100

120

140

ΣsPCAFlux (µg m-2y-1) 0

20

40

60

80

Production (kt) Fig. 9. Lake DV09 short-chain polychlorinated n-alkane (AsPCA) organic carbon and focus corrected flux, the historical diatom concentration trend (Outridge et al., in press), and the estimated sPCA global production (Muir et al., 2000b).

G.A. Stern et al. / Science of the Total Environment 342 (2005) 223–243

Diatoms (cells/g DW)

A 1e+5 2000

1e+6

1e+7

1e+8

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Year

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1920 1900

0

5

10

15

20

25

Flux (ng m-2y-1)

B 1997

1978

1963

km), whereas Lake DV09 is over 4000 km from sources in the eastern United States. In the 1950s and 1960s, sediment core slices were dominated by OCDD (Fig. 10B), suggesting that PeCP was the primary PCDD/F source. From 1970 to 1993, the TCDFs were the most abundant homologue group, indicative of combustion sources. OCDD reasserts its dominance in the surface slice. This may be due to improvements in combustion technology that have reduced PCDD/F emissions from commercial incinerators (Alcock and Jones, 1996), reducing the proportion of TCDF relative to OCDD. However, the PCDD/F homologue profile in the surface sediment does not resemble those recently measured in air from Alert, where the more volatile PCDF congeners predominate (Hung et al., 2002b). Commoner et al. (2000), using the HYSPLIT (Hybrid Single-particle Lagrangian Integrated Trajectory) air transport model, determined the marine flux of PCDD/F at Arctic Bay (the nearest model site to Lake DV09) to be 7.9 pg TEQ m
2 y
1. This value is approximately 9-fold higher than the surface flux of 0.9 pg TEQ m
2 y
1 calculated for DV09, but is similar to the maximum PCDD/F flux of 8.7 pg TEQ m
2 y
1, measured in the slice dated to 1978 (Table 1). Arctic Bay is approximately 250 km southeast of Lake DV09, and the difference between estimated and measured flux may be partly attributed to the greater distance of Lake DV09 from PCDD/F sources. However, Commoner et al. (2000) pointed out that it is more appropriate to use models such as HYSPLIT to estimate relative rather than absolute rates of deposition and to determine sources. Absolute deposition rates can be far more effectively determined from actual measurements of environmental samples. 3.2.9. Polybrominated diphenyl ethers (PBDEs) Polybrominated diphenyl ethers (PBDEs) are used as flame retardants in materials such as electronic equipment, paint, textiles, and polyurethane foam in

1951

4F 5F 6F 7F 8F

30

239

4D 5D 6D 7D 8D

Fig. 10. (A) Lake DV09 APCDD/F organic carbon and focus corrected flux; (B) PCDD/F homologue profiles in a selection of sediment slices (numbers refer to the number of chlorines; D=polychlorinated dibenzo-p-dioxin; F=polychlorinated dibenzofuran; scaled to most abundant congener=100%).

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upholstered furniture. Three commercial PBDE formulations, referred to as penta, octa, and deca, have been produced by various manufacturers since the late 1970s (de Boer et al., 2000). Compounds from these mixtures, especially the penta formulation, have been detected in many environmental compartments, including arctic marine mammals (Stern and Ikonomou, 2001; Ikonomou et al., 2002; Law et al., 2003), fish (Stern et al., 2001), marine sediments (Stern and Evans, 2003), and air (Alaee et al., 2003). As a result, the European Union (EU) recently adopted restrictions on the marketing and use of both the penta and octa PBDE mixtures. APBDE concentration in the DV09 surface sediment was 0.25 ng g
1 dry weight (Table 1), similar to the values reported by Stern and Evans (2003) for marine surface sediments in the Canadian arctic. As was also observed in the marine sediments (Stern and Evans, 2003), BDE 47 predominates in the DV09 sediment and together with BDE 99 and BDE 100 accounts for over 80% of APBDE. Maximum PBDE flux was at the sediment surface, although PBDEs were detected in all sediment slices, even after blank correction. This unusual finding is probably not due to chemical diffusion within the sediments, as it is not apparent in any of the other chemical classes, most of which have lower K OW values than the PBDEs and would be expected to be more mobile. It may instead be due to difficulties in PBDE analysis. Because PBDEs are ubiquitous in modern furniture and buildings, their presence is difficult to control. A surface maximum is consistent with PBDE production, as others have found increasing PBDE concentrations over the last few decades in both arctic marine mammals (Ikonomou et al., 2002) and European sediments (Zegers et al., 2003). However, as was observed with many of the other compound classes discussed above, the near-surface PBDE maxima in DV09 might also be attributed to changes in primary productivity and/or anaerobic degradation of PBDEs in older slices.

than most arctic lake sediment cores, many of the HOCs in Lake DV09 sediments still exhibited maxima at or near the sediment surface. Current assumptions are that transport processes of HOCs from atmosphere to sediments via the catchment and water column have not changed in recent history and, therefore, that the trends in sedimentary HOC flux values simply and directly reflect trends in the atmospheric concentrations of the various HOCs and possible anaerobic degradation as noted above. However, the authors postulate that the substantial increases in algae productivity which seems to coincide with the recent cycle of the Arctic Oscillation (Macdonald et al., 2004), a decadalscale climatic phenomenon which has resulted in warmer temperatures in much of the Arctic, and which appears to be superimposed over an overall long-term warming trend (Macdonald et al., 2005; Serreze et al., 2000), may at least be partially responsible for the increased HOC flux observed since the early 1980s. The possible effects of climate warming on HOC fluxes in high Arctic Lakes as well as in Marine systems such as the Beaufort Sea and Hudson Bay require further study as one potential result may be to once again increase or at least to maintain current HOC concentrations in some food chains which support traditional foods important to northerners (Macdonald et al., 2005).

4. Conclusions

References

In general, HOC concentrations and inventories in Lake DV09 sediments were similar to those of other Arctic lakes but, despite a higher temporal resolution

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Acknowledgements We thank R.W. Macdonald for his insightful discussions and suggestions, T. Siferd for his assistance with core collection, and two anonymous reviewers for their comments. Financial support was provided by the Environmental Science Strategic Research Fund (Fisheries and Oceans Canada) and the Northern Contaminants Program (Indian and Northern Affairs Canada). Logistical support was provided by the Polar Continental Shelf Project.

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