The rates of accumulation and chronologies of atmospherically derived pollutants in Arctic Alaska, USA

the Scieyx of the TotalEilwmmment A”l-,aaul~r-ll- l”mUrE”-dlbR-p”nb~

The Science of the Total Environment 160/161(1995) 347-361

The rates of accumulation and chronologies of atmospherically derived pollutants in Arctic Alaska, USA

C.P. Gubala”“, D.H. Landersa, M. Monettib, M. Heitb, T. Wade”, B. Lasorsad, S. Allen-Gil”

aU.S. EPA Entironmental Research Laboratory, 200 SW35th St., Coruallis, OR 97333, USA

bU.S. Department of Energy, New York, NY, USA

‘Geochemical Environmental Research Group, College Station, TX, USA dBattelle Pacific Northwest Laboratories, Sequim, WA, USA

eOregon State Uniwrsi& Cord& OR 97331, USA

Abstract

Anthropogenically derived pollutants (e.g. trace metals, organochlorines, radionuclides) are deposited upon arctic ecosystems. To determine the range of probable biotic effects of these pollutants, one must know the rate at which they enter and are retained within an ecosystem. However, unknown deposition mechanisms and the complexity of quantifying atmospheric concentrations of constituents of interest make direct measurements of pollutant flux to arctic terrestrial and aquatic ecosystems difficult and/or impractical. Methods of indirectly measuring rates of pollutant accumulation, such as lake sediment stratigraphic analyses, can fill this void. Present in the sediment of two Alaskan lakes sampled in April 1991 were quantifiable concentrations of numerous organochlorine compounds, including DDT and its metabolites (Sum, 0.05-0.60 rig/g), PCBs (Sum, 0.20-30 rig/g), and lindane (0.20-0.80 rig/g). These surface concentrations correspond to estimated deposition rates of 2-6 ng/m’/year (XDDT + metabolites), lo-300 ng/m’/year (ZPCBs), and g-10 ng/m*/year (lindane). The rates and chronologies of accumulation of these pollutants and others are discussed with regard to the process of long-range atmospheric transport.

Keywords: Accumulation; Pollutants; Alaska

1. Introduction

Over the past two decades, researchers have detected the presence of anthropogenic pollutants in a variety of environmental ‘compart- ments’ in the Arctic (Barrie, 1986; Ottar, 1989; Shaw and Khalil, 1989; Welch et al., 1991). Noted

* Corresponding author.

pollutants include radionuclides, trace metals, acid precursors, and chlorinated hydrocarbons (Hargrave et al., 1988,1989; Gregor and Gummer, 1989; Muir et al., 1992). The presence of even small quantities of organic constituents within arctic ecosystems is of concern, due to their potential to biomagnify efficiently through the lipid-based food webs (Kinloch et al., 1992).

Recently, much effort has been devoted to documenting the occurrence of anthropogenic

0048-9697/95/$09.50 0 1995 Elsevier Science BV. All rights reserved. SSDI 0048-9697(95)04368-W

348 C.P. Gubah et al. /Sci. Total Enuion. 160/161 (1995) 347-361

pollutants in arctic regions and describing potential deposition venues and ecosystem effects (Bar- rie et al., 1992). However, a few researchers have attempted to measure the rate of accumulation of specific pollutants or to document their historical pattern of deposition (Cornwall, 1986; Strachan, 1988; Johansson, 1989; Hermanson, 1990; Mu- droch et al., 1992). These deficiencies derive largely from a combination of the limited access and logistic complexities posed by the study region. However, measurement of the chronologies and accumulation rates of anthropogenic pollutants are quite important when considering (1) the relationship between the chronology of pollutant deposition and the rate of emission from probable sources, (2) the relationships between specific rates of pollutant accumulation and rates of ecosystem response, and (3) projections of fu- ture patterns and rates of pollutant deposition.

Objectives of an existing U.S. EPA program, the Arctic Contaminant Research Program (ACRP; Landers et al., 19921, are to differentiate between natural and anthropogenic sources of contaminants to the U.S. Arctic, estimate the timing and onset of contaminant loading, and establish current concentrations of contaminants in arctic ecosystems. The ACRP also seeks to determine the potential biological effects of atmospherically derived pollutants upon aquatic ecosystems. These tasks are accomplished par- tially through the collection and analysis of arctic lake sediments. Natural and anthropogenic pollutant sources are differentiated within lake sediments by comparing recent concentrations and accumulation rates with historical or background conditions. In this way, a discernible change in the depositional pattern of a pollutant during recent times may signify the onset of an anthropogenic contamination signal.

The accumulation of anthropogenic pollutants in the sediments of an arctic (Schrader Lake) and a subarctic (Wonder Lake) Alaskan lake are compared in this manuscript. This spatial and temporal comparison yields information relating to the possible transport pathway of airborne pollutants. The rates of accumulation of specific pollutants are presented and compared between sites.

2. Methods

Two Alaskan lakes, Wonder Lake (63”28’N, 150”52’W, Denali National Park and Preserve) and Schrader Lake (69”22’N, 144”6O’W, Arctic National Wildlife Refuge) were selected as the first sediment study sites of the ACRP, based upon the availability of primaly limnological data for these systems (Fig. 1). Both are large, glacial lakes with stable bathymetries and boundaries and presumably continuous and regular sediment accumulation rates through at least the past 150 years (Wilding, 1940; Hobbie, 1961, 1962; Werner et al., 1990). The systems are relatively remote, with limited and controlled access. These at- tributes usually contribute to minimize the anthropogenic ‘noise’ of watershed disturbance re- flected within lake sediments. But there is little evidence to support or refute this claim and the results presented in this manuscript should be cast in this light. However, if these lakes have been relatively undisturbed by local forces, then it may be possible to resolve the accumulation of atmospherically derived pollutants from their sediment stratigraphies.

Sediment cores were retrieved from Wonder and Schrader Lakes during the late winters of 1991 and 1992, respectively, using a 12.7-cm diameter gravity style corer lowered through holes cut in the ice. The sediment was processed imme- diately on site.

Initial visual determinations of core quality were made in the field. Only cores that appeared undisturbed and true (e.g. level) were subsectioned. The cores were subsectioned at l-cm intervals from 0 to 10 cm (yielding approximately 125 cc sediment per section), and at 2-cm intervals from 20 to 40 cm or the maximum depth of the core (yielding approximately 250 cc sediment). Subsectioning of cores at a finer increment than 1 cm was unacceptable, since it was necessary to process at least 100 cc of wet sediment for analysis of trace quantities of organic analytes. Sedi- ment samples were stored cold (approximately 4°C) until the time of analysis.

Sediment intervals were analyzed sequentially for total water and carbon content, radionuclides, chlorinated hydrocarbons, and metals. Radiomet-

C.P. Gubala et al. /Sci. Total Environ. 160/161 (1995) 347-361 349

0 Sediment history

0 200 I I

kilomelerr

Alimulhol tquidirtont projection

Arctic Contaminants Reseotch Program

350 C.P. Gubala et al. /Sci. Total En&on. I60/161 (1995) 347-361

ric analyses of ‘lOPb, 137Cs, and 7Be were first utilized in a constant rate of supply (CRS) (Rob- bins, 1978) model to assign approximate dates of deposition to individual sediment intervals. Based upon this information, an analytical strategy was structured for chlorinated hydrocarbons and metals for each case.

Total organic carbon (TOC) was determined by digestion of an aliquot of sediment followed by infrared absorption spectrophotometry. The TOC values given here represent a direct measurement of organic carbon rather than a surrogate, such as loss on ignition CLOD, which is more typical of sediment studies.

Radionuclide analyses were performed using y-spectrometric methods similar to those described by Gogolak and Miller (1977) and by the U.S. DOE Environmental Monitoring Laboratory (EML, 1992). Wet or freeze-dried sediment samples were packed into containers of a fixed geometry and sealed at least 21 days prior to analysis to establish secular equilibrium between 226Ra and its gamma-emitting daughter products c214Pb and 214Bi). Samples were counted on either a high-resolution germanium well detector or n-type coaxial germanium detector, depending on the type of sample container utilized. Count times varied from 1 to 4 days; longer counting times were used to improve counting statistics of lower activity samples. The quantification of radionuclide peaks in the spectrum was obtained through standard computer integration software.

Assessment of chlorinated hydrocarbons in lake sediments within arctic Alaska required their measurement to concentrations as low as 0.15 rig/g dry weight sediment in the presence of interfering sediment constituents (e.g. sulfur compounds). From 4 to 10 grams dry weight equiva- lents of sediment were Soxhlet extracted with methylene chloride and concentrated. Since the amount of water in the sediments ranged from 44 to 96%, as much as 100 g of wet sediment was extracted. Silica gel/alumina cleanup with acti- vated carbon was used to purify the sample prior

to analysis. The supernatant was also acidified to eliminate interference from carbon-bonded sulfur compounds.

Chlorinated hydrocarbons were quantified in the extracted supernatant using high-resolution gas chromatography (GC) and electron capture detection (ECD). The GC was equipped with a 30-m X 0.25mm i.d. fused silica capillary column with DB-5 bonded phase. The quantification of analytes in internal standard solutions, matrix recovery spikes, and standard reference materials were used to provide calibration and quality assurance for the GC method. The concentrations of polychlorinated byphenyls and chlorinated pesticides were determined and reported as rig/g dry weight sediment.

To prepare arctic lake sediment for trace and major metal analyses, a l- to 5-g aliquot was totally dissolved in a sealed Teflon pressure vessel using a two-step digestion with high purity nitric, perchloric, and hydrofluoric acids. Quantification of metals in the extract was then affected by use of inductively coupled plasma mass spectrometry (ICP/MS), graphite furnace and flame atomic absorption, energy dispersive X-ray fluorescence (XRF), and cold vapor atomic absorption (CVAA). The quantification of analytes in internal standard solutions, matrix recovery spikes, and standard reference materials were used to provide calibration and quality assurance for each analytical method.

Replicate analyses were conducted on 10% of the samples to assess precision of the organic and inorganic analytical techniques. Duplicate values of the concentrations of analytes reported in this manuscript varied by no more than 5%.

3. Results and discussion

3.1. Sediment characteristics and rates of accumulation

The physical characteristics of sediment from Schrader and Wonder Lakes differed substan- tially (Fig. 2; Table 1). The percentage of water

Fig. 1. Map of Alaska, USA, depicting the location of Wonder and Schrader Lakes and their geographic orientation within the prominent topography.

C.P. Gubala et al. / Sci. Total En&on. 160 / 161(1!995) 347-361 351

Schrader Lake

5 Q -30 n

-40

-50 40 65

H20 % TOC % 2 10Pb Bq/gdw 137Cs Bq/gdw

0

-5

-10 r-l

-15

-20

-25 70 100

H20 %

Sed. Act. Rt. w/mVyr

Wonder Lake

TOC % 210Pb Bq/gdw 13i’Cs Bq/gdw Sed. Act. Rt. m4/m2/yr

12

Fig. 2. Percent water (%H,O), total organic carbon (%TOC), unsupported ‘lOPb and ‘37Cs, and CRS sediment accumulation rates versus depth in the sediment for Wonder and Schrader Lakes.

near the sediment surface of Wonder Lake was larly to 73% and 3.5%, respectively. This change 96% and contained low concentrations of organic is consistent with a pattern of diagenesis combin- carbon (TOC, 7.3%) relative to lakes of lower ing microbial processing and compaction of the latitudes. Deeper in the sediment profile of Won- sediment. By contrast, Schrader Lake surface der Lake, the %H,O and %TOC decreased regu- sediments are extraordinarily low in %H,O and

Table 1 Site and sediment characteristics of Wonder and Schrader Lakes

Characteristics Wonder Lake Schrader Lake -

Site Location Maximum depth (m) Lake area (km*) Watershed/lake area (km*) Coring date

Sediment % Water % Total organic carbon

63”28’N, 150=‘52’W 69”22’N, 144”6O’W 76 57 3 13

10 10 May 1991 May 1992

13-96 44-61 3.5-7.3 0.5-1.2

3.52 C.P. Gubala et al. /Sci. Total En&on. 160/161 (1995) 347-361

%TOC (77% and 1.2%, respectively). Percent H,O shows an erratic pattern with depth in the Schrader sediments, whereas TOC remains relatively constant at about 1%.

Mixing of the upper layers of the sediment for both lakes appears to be minimal for several reasons. Unsupported “‘Pb profiles appeared to decrease in a regular exponential manner from the surface. And since TOC is very low in the subsurface sediments from both lakes, utilization of carbon in these systems is likely to be limited to a thin sediment/water interface region. 7Be, a naturally occurring, short-lived radionuclide, was undetected in all samples and hence could not be used to infer a mixing depth.

The distribution of radionuclides in the sediment from the two coring sites in Wonder and Schrader Lakes are relatively similar. The entire inventories of 137Cs and unsupported “‘Pb were found in the top 5-7 cm of the cores (Fig. 2). Unsupported “‘Pb decays within roughly 140-150 years and its absence below approximately 5 cm in each core indicates relatively slow sediment accumulation rates for both lakes. Two methods were used to calculate the average rate of sediment accumulation for these two lakes during recent times. A CRS model (Robbins, 1978) was used to determine the sediment accumulation rate on an interval specific basis. This calculation yielded accumulation rates averaged over the thickness of a core slice (1 cm). For comparison and verification, a simple average (SA) accumulation rate was derived by dividing the mass of sediment accumulated above the point where unsupported 210Pb reaches zero by 150 years. Using these methods, the average rate of sediment accumulation over the past 150 years is approximately 11 and 55 g/m2/year for the Wonder and Schrader cores, respectively (Fig. 2). Further, since the calculated rate of sediment accumulation is nearly constant within each core (within +5%), units of concentration rather than flux are used to describe the pollutant distribution (Figs. 2-9). There is no difference between the observed concentration or flux patterns of pollutants in each core, since in the instance of a constant sediment accumulation rate, the two vary by a simple scaler.

Due to the extremely low sediment accumulation rates in both lakes and presumably limited mixing from bioturbation, CRS and SA calcula- tions seem appropriate. This approach is further substantiated when the results of the radionuclide analyses are compared to the general sampling procedures. For the large diameter core (12.7 cm), required to collect enough sample for organic analyses, slices 1 cm thick were considered to be the thinnest practical increment of subsectioning. All the “‘Pb record was then retrieved in just 5-7 samples, reducing the statistical power of log-linear based sedimentation models. And since sediment mixing seems to be restricted to the extreme surface of both cores, applying a detailed mixing model to the small number of samples comprising the ‘i”Pb record is also unnecessary.

Differences in the mass accumulation rates for the two systems resulted largely from the relative density differences of the two sets of sediment. An active glacier resides within the Schrader Lake watershed and is probably responsible for the lower amount of carbon and relatively higher density of Schrader’s sediments.

The top centimeter of sediment for each core may have taken up to 30-40 years to accumulate. After accounting for diffusion and diagenesis, we would expect to find evidence for anthropogenic pollution as enrichments of specific constituents within the top l-5 cm of sediment. For trace metals, increasing concentrations from a depth of 5 cm to the surface may indicate anthropogeni- tally derived contamination. Since halogenated hydrocarbons have been manufactured only recently (I 50 years), effects from these compounds should be noted only in the surface sediments of both lakes and can all be attributed to anthropogenic inputs.

3.2. Focusing factors The aforementioned linear and mass sediment

accumulation rates are specific to the individual cores taken from each lake. In deep lake basins, sediment typically accumulates at a rate faster than the lakewide average due to the physical process of sediment focusing (Davis and Ford, 1982). To determine the focusing factor for each core, the total inventories of unsupported ‘i”Pb

C.P. Gubala et al. / Sci. Total En&on. 160 / 161 (1995) 347-361 353

and 137Cs in the core (M. Monetti, Department of Energy, personal communication, 1993) were compared to an estimate of the total inventories of the same radionuclides for the surrounding terrestrial soil region. The latter can be determined directly through unit surface area soil measurements or by interpolating from proximate atmospheric collectors. Using a combination of these techniques, it was determined that the sediment focusing factors for the Wonder and Schrader Lake cores were approximately 1.4 and 1.5, respectively. A focusing factor of 1.5 for a specific core indicates that sediment accumulated at a rate 50% faster in that core than in the lake as a whole. These focusing factors refer specifi- cally to the individual cores taken at Wonder and Schrader Lakes.

Normalizing the linear and mass accumulation rates for the cores from each lake by the core- specific focusing factors leads to an estimate of the average rates of sediment accumulation for each lakewide system. The average lakewide mass accumulation rates for Wonder and Schrader Lakes were 11 and 55 g/m2/year, respectively. These figures are the most relevant for expressing and understanding the accumulation of pollutants in these lakes, as it is vital to normalize the observed concentrations of pollutants in the sediments by a realistic rate of average sediment accumulation.

3.3. Metals The patterns of trace and common metals found

in sediments must be interpreted relative to their crustal abundance and potential to change redox states. The latter is quite important, as microbial processing of organic carbon comprising a fraction of the sediment typically creates an anaerobic environment within the sediment and inter- stitial porewater. Metals bound in oxy-hydroxide solid phases (e.g. Fe, Mn) under the aerobic conditions of the lake and watershed may be liber- ated within the sediment to migrate vertically along a redox gradient. Therefore, surface enrichments of redox sensitive elements do not neces- sarily represent anthropogenic input, but may rather reflect natural diagenesis.

In Wonder and Schrader Lakes, trends in Fe

and Mn demonstrate a redox-driven diagenetic phenomenon (Figs. 3,4). A large fraction of the total concentrations of these elements are typically deposited as Fe(III)- and Mn(IV)-based solid phases. Once buried to a depth of high electron density, resulting from organic carbon bioprocess- ing, Fe(II1) and Mn(IV) solid phases are dissolved and the elements subsequently reduced to Fe(H) and M&I). These elements then move upward along vertical diffusion gradients and are reoxi- dized and precipitated at the zone of notable change in redox conditions in the sediment profile. The profiles of Fe and Mn solid phases in Wonder Lake are similar to those found in other lakes, with their diagenetic enrichments occurring near the sediment-water interface (Cornwall, 1986; Gubala and White, 1987; White et al., 1989). Iron and Mn in Schrader Lake differ markedly, however, as the enrichment of these elements occurs nearly 20 cm down into the core. These atypical profiles probably result indirectly from the low amount of organic carbon in the sediment. The zone of notable redox change for the Schrader sediment must occur below 20 cm, suggesting that the top 20 cm of sediment exists in an aerobic environment. This is significant when later considering the aerobic versus anaerobic pathways for the redox reactions of other pollutants. In particular, a deep zone of redox change in the sediment may inhibit methylation of important biotoxicants, such as Hg (Stumm and Morgan, 1981). Finally, while Fe and Mn comprise the largest mass migration of metals in the sediment, As follows a similar pattern for Schrader Lake (Wonder Lake As data not available). This is significant, as As is a biotoxicant and may form a methylated compound that is readily available for biological uptake (Stumm and Morgan, 1981). Since the bulk of this element is trapped below 20 cm of sediment and methylation near the sediment-water interface is limited, its significance to aquatic biota is diminished.

To assess anthropogenic enrichment of nonre- dox sensitive metals in the Arctic (e.g. Cu, Zn, Ni, Pb, Hg, V> (Amundsen et al., 19921, it is important to characterize the watershed processes that may lead to natural enrichments. Profiles of the metals that may be considered as pollutants are

354 C.P. Gubah et al. /Sk. Total Emiron. 160/161 (1995) 347-361

a

B

2 8

-42 80 I

I I 30 loo

[F4 mddr 120

O.PYl -I ‘.l!

[WI Wtidw

1 40 70

[Cd urlhd”

30

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Schmder L&c

Fig. 3. Concentrations of major and trace metals versus depth in the sediment of Schrader Lake.

’ El 70 110

[VI w&w WI m&h@*

“a 8

-2s H 100 250

wonder lake

Fig. 4. Concentrations of major and trace metals versus depth in the sediment of Wonder Lake.


compared to the aluminium profile, since sediment profiles of Al in the U.S. Arctic reflect the change in erosional inputs from watersheds (Cornwall, 1986). Schrader Lake exhibits an Al profile that indicates a slight but steady increase in erosional inputs over several hundred years, whereas the Wonder Lake profile shows a near 40% decrease in the rate of elastic inputs over just about 80 years (Fig. 3). It is beyond the scope of this paper to investigate the causes of these trends. Rather, it is important to interpret the trends in metals that are potential candidates as anthropogenically derived pollutants relative to the trends in elastic inputs represented by Al.

Al, suggesting that their presence is due largely to natural phenomena.

Precise and accurate quantification of enrichment factors of Pb, Hg, and V in Wonder Lake by a direct element-to-aluminum ratio is not entirely defensible (Figs. 3,4). Although Al provides a useful benchmark for this comparison, not enough is known about the site-specific watershed processes that affect the relationships between modes of transport and deposition of different elements into the lake. Therefore, the enrichment values of these elements will be considered as estimates in subsequent discussions (Table 2).

In Schrader Lake, none of the metals typically implicated as long-range atmospheric pollutants increase at a rate faster than that of Al (Fig. 3). In addition, the degree of variance of most metals with depth in the sediment of Schrader Lake limits the ability to detect slight increases near the surface. Within Wonder Lake, however, there exists strong evidence for a recent enrichment in Pb relative to elastic inputs (Fig. 4). The sediment record also suggests a very recent enrichment in V, but the evidence is much less compelling. Other elements (Zn, Cu, Ni) mimic the pattern of

3.4. Organochlorine compounds The most important observation regarding

organochlorine compounds in the sediments of Schrader and Wonder Lakes is that they are present (Figs. 5-9; Table 2). Since these compounds have no natural sources and were presumably not applied directly into the lakes or watersheds, their presence alone provides evidence to support the hypothesis of transport of these constituents to the lakes via the atmosphere from non-local sources. The closest regional sources of organochlorines are from developed

Table 2 Concentrations and recent watershed accumulation rates of anthropogenic pollutants in Wonder and Schrader Lakes”

Contaminant Wonder Lake Schrader Lake

Surface sediment enrichment Total HCHs (rig/g) Total chlordanes (rig/g) Total DDTs (rig/g) Total PCBs (rig/g> Hg ( &d Pb ( a/p> v ( M/g>

0.810 0.402 0.569

30.706 0.03b

12.00b 30.00b

0.193 0.035 0.051 0.172 No enrichment No enrichment No enrichment

Recent watershed accumulation rate Total HCHs (ng/m*/year) 6.4 7.1 Total chlordanes (ng/m*/year) 3.1 1.3 Total DDTs (ng/m’/year) 4.5 1.9 Total PCBs (ng/m’/year) 241.2 6.3 Hg ( kg/m’/year) 0.2 No enrichment Pb ( pg/m’/year) 95.3 No enrichment V ( pg/m’/year> 235.7 No enrichment

aAnthropogenic values listed for inorganic materials were determined by subtracting out the natural background. Watershed accumulation rates are corrected for the focusing factors for each sediment core: (watershed accumulation rates) = (surface sediment enrichment) X (surface sediment accumulation rate)/(focusing factor).

bEstimated value.

356 C.P. Gubah et al. /Sci. Total Entiron. 160/161 (1995) 347-361

0.8

0.6 2 B

2 P 0.4

\”

2

0.2

Surfacs Sedimenb (0-4cm)

Fig. 5. Concentrations of 1245 TCB, 1234 TCB, PC benzene, PC anisole and HCB in the surface sediments of Wonder and Schrader Lakes. Note that nearly the entire inventories of the analytes were found within the top 1 cm of sediment.

areas in Alaska such as Fairbanks and An- chorage, where these compounds might have been used for commercial purposes. Wonder Lake is much closer than Schrader Lake to these possible Alaskan sources.

All classes of organochlorine compounds show similar distributions with sediment depth within each core. In both Wonder and Schrader Lakes, the entire inventories of organochlorine com-

Surface Sediment8 (O-4cm)

1

Fig. 6. Concentrations of HCH isomers including yHCH Fig. 8. Concentrations of heptachlors and chlordanes in the (lindane) in the surface sediments of Wonder and Schrader surface sediments of Wonder and Schrader Lakes. Note that Lakes. Note that nearly the entire inventories of the analytes nearly the entire inventories of the analytes were found within were found within the top 1 cm of sediment. the top 1 cm of sediment.

0.4 t

r I

0.3

d t

2 v 0.2

2

2

0.1

%d i B -?J /bJ

Fig. 7. Concentrations of DDT and its metabohtes in the surface sediments of Wonder and Schrader Lakes. Note that nearly the entire inventories of the analytes were found within the top 1 cm of sediment.

pounds were found within the top 4 cm of sediment, with approximately 95-98% of the mass occurring in the top 1 cm. The top 4 cm of both lakes were analyzed and presented for a uniform comparison of the total organochlorine pool in both lakes. Assuming that diffusion, bioturbation, and/or core smearing accounts for the 2-5% of

0.20

0.15

s sr

4 0.10

\”

2 0.06

0.00

Surface SedlmenLs (O-4cm)

C.P. Gubala et al. / Sci. Total Environ. 160/ 161 (1995) 347-361

0.2 Surface Sediments (O-4cm)

5

2 -d 0.1

T!

2

0.0

4.0-

351

PCB Congeners #

Fig. 9. Concentrations of PCB congeners the surface sediments of Wonder and Schrader Lakes. Note that nearly the entire inventories of the analytes were found within the top 1 cm of sediment.

material found from 1 to 4 cm in sediment depth in each lake, this distribution indicates that these anthropogenic compounds impacted both lakes sometime over approximately the last 40 years. While this observation should not be surprising considering the production timing of these compounds, the distribution patterns indicate that bioturbation of surface sediments is negligible and that the coring procedure itself did not intro- duce any significant error from translocation or contamination.

The concentrations of organochlorine constituents varied between the two lakes with some consistency. In general, the surface sediments of Wonder Lake were contaminated by organochlorines at concentrations 5-10 times higher than those of Schrader Lake. In particular, DDT and its degradation products, chlordane congeners, y-HCH (lindane), PC anisole, and HC benzene were all present in readily quantifiable amounts in Wonder Lake. The same compounds were barely detectable in Schrader Lake.

However, when expressed as fluxes, normalized

by sediment accumulation rates, the comparison of organochlorines between lakes changes (Table 2). Since the mass flux at Schrader Lake was five times higher than at Wonder Lake, the lower concentrations of organochlorines at Schrader Lake translate into pollutant fluxes quite similar to those at Wonder Lake.

The concentration of the total PCB congeners varied quite differently between lakes. In Schrader Lake, the concentration of the sum of PCBs seemed to be in accord with the concentrations of other classes of organochlorine compounds, suggesting a systematic mechanism for the introduction of these compounds into the lake. It is generally believed that this mechanism is long-range atmospheric transport, since the relative concentrations of organochlorines in Schrader Lake appear to fall within a range comparable to their relative differences in vapor pressure.

In Wonder Lake, however, the concentration of ZPCBs was higher than other organochlorine compounds in the sediment. While Wonder Lake is generally protected from immediate anthro-

358 C.P. Gubala et al. /Sci TotalEntiron. 160/161 (1995) 347-361

pogenic activity, due to its location within a national park, it seems likely that the PCB ‘signal’ in this lake may have come from direct contamination rather than from long-range atmospheric transport. Supporting this ‘contamination event’ hypothesis is the disparity between the behavior of PCBs relative to the other semi-volatile organochlorines with similar physical/chemical characteristics. While concentrations of total HCHs, chlordanes, and DDTs were systematically higher by a factor of 2-6 in Wonder Lake over Schrader Lake, PCBs were enriched loo-fold. Differential atmospheric transport among all of these semi-volatile compounds varies with their vapor pressures. But the PCB enrichment in Wonder Lake falls outside of a logical range of variance relative to the other compounds.

However, due to the relatively low temporal resolution derived from the sediment stratigraphy, it is impossible to directly confirm or reject the hypothesis of a contamination ‘event’ affect- ing Wonder Lake. Similarly, though, it cannot be assumed with confidence that PCBs were introduced into Wonder Lake by long-range atmospheric transport alone.

3.5. Mass accumulation of anthropogenic pollutants All organochlorine compounds found in both

lakes were derived from anthropogenic activity. Of these compounds, it is assumed that the HCHs, chlordanes, and DDTs were derived from atmospheric sources, since there are no other likely local sources of these compounds that affected either watershed. Schrader Lake is completely remote with very little human visitation. And while anthropogenic activity near Wonder Lake supports tourism within the surrounding national park, there is no record of the direct application of HCHs, chlordanes, or DDTs at this site. An accidental release of PCBs at Wonder Lake is a possibility, however, since electrical and hydraulic equipment has been used within the watershed. And as was previously discussed, PCB enrichment in Wonder Lake greatly exceeds the level of enrichment of other organochlorines that should behave similarly in the atmosphere. Part of the total PCBs found in Wonder Lake may indeed be due to atmospheric deposition, but cannot be

separated from the total signal based upon the available data. However, if the PCBs follow the same patterns of deposition as the other organocholorine compounds (e.g. concentrated 2-6 times higher than in Schrader Lake), then as much as 10% of the 30 rig/g in Wonder Lake may be due to atmospheric sources.

Of the metals, large fractions of the Pb (50%), and small amounts of Hg (10%) and perhaps V (10%) in the sediment can be considered of anthropogenic origin in Wonder Lake. No quantifiable amounts of the metals in the sediment of Schrader Lake have resulted from anthropogenic activity. Table 2 presents the surface sediment concentrations and rates of accumulation of major classes of organochlorine compounds and of excess Hg, Pb, and V in the sediments and watersheds of the two lakes. The area1 mass accumulation rates of these pollutants across an entire watershed are assumed to be equal to the mass accumulation rates of each core, after correcting for the focusing factor. This presumes that a pollutant is deposited upon the lake and its watershed in a uniform manner. These values are akin to values derived from direct atmospheric collectors and represent an estimate of the rates of pollutant transfer from the atmosphere to the terra.

The calculated watershed accumulation rates for organochlorines at both sites are on the order of l-10 ng/m2/year. The exceptions to these rates are those of PCBs in Wonder Lake, which appear to have been introduced by mechanisms other than atmospheric deposition.

Of the compounds that appear to have been deposited via atmospheric transport, the two lakes have received about the same amount of organochlorines on a unit area basis (Table 2). Of the classes of compounds investigated, the HCHs were found to accumulate in both systems at a rate of as much as 2-5 times greater than chlordanes or DDTs. Since HCHs are more volatile than the latter compounds, it has been hypothe- sized that their transport into the Arctic from lower latitudes would occur faster than that of heavier compounds (Wania and Mackay, 1993). This ‘global distillation’ hypothesis predicts that high latitude regions initially receive a lower rate

C.P. Gubala et al. /Sci. Total En&on. HO/161 (1995) 347-361 359

of deposition of heavier organochlorines than lower latitude regions as the compounds propa- gate slowly northward along a theoretical atmospheric temperature gradient. Consistent with this hypothesis, the more northerly Schrader Lake receives a higher proportion of lighter, more volatile compounds (HCHS) over heavier compounds (chlordanes, DDTs), since it is at a greater distance from lower latitude source areas. Pro- vided that the data obtained from Wonder and Schrader Lakes reflect long-range atmospheric transport processes, these results support the global distillation theory.

Interpretation of the rates of accumulation of PCBs is confounded by the extraordinarily high concentrations found in Wonder Lake. In Schrader Lake, however, the least chlorinated, most volatile PCBs dominate, which is consistent with the results from the other organochlorines (Fig. 9).

While the interpretation of the organochlorine observations seem consistent with a logical theory, great care must be taken to assign the proper weight to these data. The concentrations of the constituents were extremely low and the relative uncertainty behind the calculation of sediment accumulation rates may be as high as 20-30%. It is then conceivable that the propagated sampling and analysis error could account for the difference in accumulation rates amongst compounds and between lakes.

Anthropogenically derived Pb accumulated in Wonder Lake at a much higher rate than in Schrader (Figs. 3,4, Table 2). Indeed, due to the high background concentrations of metals in the sediment of both lakes, accumulation rates of anthropogenically derived Hg, Pb and V could be calculated only for Wonder Lake. And of these elements, only Pb in Wonder Lake showed sufficient systematic enrichment to be ascribed with certainty to anthropogenic causes. Lead from anthropogenic sources has recently accumulated in the Wonder Lake watershed at a rate of approximately 100 pg/m’/year. This rate of enrichment constitutes approximately a twofold increase in the rate of accumulation from natural sources. The anthropogenic sources causing Pb enrichment in Wonder Lake are unknown, but a likely

source is usage of leaded gasoline fuel in North America (Valette-Silver, 1992). Since Wonder Lake is within a national park that holds a small road system, some of the Pb enrichment may be attributed to nearby sources. However, the available evidence is not sufficient to discern the relative contributions of Pb to Wonder Lake from other local (e.g. mining operations) or long-range atmospheric sources.

Increases in the rates of accumulation of Hg and V presumably due to anthropogenic activity are < 10% of the natural background. At these concentrations, the ecological significance of their increased rate of input to the aquatic system is perhaps minimal.

Even though the presence of an anthropogenic signal for metals in the sediments of both lakes could not be firmly established, it is still possible that either lake has been affected by the long- range atmospheric transport and deposition of these industrial byproducts. The evidence derived from the total metals profiles in these lake sediments is simply inadequate to describe all but the most pronounced inputs (e.g. Pb in Wonder Lake). This is due to the high concentrations of metals found in the bedrock and soils of the study regions (Gough et al., 1988). To further investigate the possibility of the deposition of anthropogenic metals to these watersheds through the use of sediment profiles, more detailed chemical specia- tion or isotopic fractionation of the metals should be considered. Sampling sites with much higher rates of organic matter deposition should also be considered. This would yield a clearer distinction between erosional and atmospheric inputs.

4. Conclusions

Lake sediment stratigraphies can be used to determine the current rates of accumulation of anthropogenic pollutants in northern and arctic ecosystems. However, due to the low rates of sediment accumulation for Wonder and Schrader Lakes, the chronologies of pollutant deposition for the two systems were resolved to within 30- to 40-year periods only. Regardless, the rates that were determined for organochlorine compounds provide important information in understanding

360 C.P. Gubala et al. /Sci. Total Entiron. 160/161 (1995) 347-361

anthropogenic affects within the Arctic. Further lake coring and comparison with pollutant information collected from other parts of the environment (e.g. soils, fish, mammals) will lead to a better understanding of the relationship between pollutant deposition in the Arctic and the probable sources.

Metals derived from long-range atmospheric transport may affect the two study sites, but due to the high background concentrations, it was impossible to isolate and describe this phenomenon. Apart from the difficulties presented from the relatively high and variant background signals of metals in each system, the analytical strategy was also limiting. Total analyses of metals in sediment groups elastic inputs together with atmospheric inputs to form a complicated stratigraphy. By using more elaborate analytical techniques, reactive atmospherically derived metals may be chemically isolated from resistant erosional inputs within the sediment stratigraphies.

Acknowledgements

We acknowledge the contributions of Kathy Hurley of ManTech, who participated in the sampling expeditions. Suzanne Pierson of ManTech supplied GIS data integration support. Deborah Coffey of ManTech assured data quality and pro- vided critical feedback on the conceptual aspects of this study. The research described in this arti- cle has been funded by the U.S. Environmental Protection Agency. This document has been pre- pared at the EPA Environmental Research Laboratory in Corvallis, OR, through contract 68-C8-0006 to ManTech Environmental Tech- nology, Inc. It has been subjected to the Agency’s peer and adminstrative review and approved for publication. Mention of trade names or commercial products does not constitute endorsement or recommendation for use.

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Documents

The rates of accumulation and chronologies of atmospherically derived pollutants in Arctic Alaska, USA