Results from the Lake Michigan Mass Balance Study: Concentrations

In this paper, we summarize the data and methods used to estimate atmospheric exchange of polychlorinated biphenyls (PCBs) and trans-nonachlor with La...
0 downloads 0 Views 142KB Size
Download PDF
Environ. Sci. Technol. 2001, 35, 278-285

Results from the Lake Michigan Mass Balance Study: Concentrations and Fluxes of Atmospheric Polychlorinated Biphenyls and trans-Nonachlor SONDRA M. MILLER,† MARK L. GREEN,‡ JOSEPH V. DEPINTO,‡ AND K E R I C . H O R N B U C K L E * ,† Department of Civil and Environmental Engineering, The University of Iowa, Iowa City, Iowa 52242, and Great Lakes Program and Department of Civil, Structural, and Environmental Engineering, State University of New York at Buffalo, 202 Jarvis Hall, Buffalo, New York 14260

In this paper, we summarize the data and methods used to estimate atmospheric exchange of polychlorinated biphenyls (PCBs) and trans-nonachlor with Lake Michigan. This work was conducted as part of the Lake Michigan Mass Balance (LMMB) study. For the atmospheric component of the LMMB, more than 400 gas- and particulate-phase samples were collected at eight sites on the shore around the lake (shoreline) and at 14 sites on the lake (overwater). We review the quality of the data set; describe the concentrations in atmospheric gas and particulate phases; report local, instantaneous, net gas fluxes; and estimate annual deposition of the particle-associated compounds. The quality of the data set is high except for a subset of overwater samples where PCB contamination is suspected. Gasphase trans-nonachlor concentrations (although not the resulting gas fluxes) are inversely correlated with latitude and positively correlated with temperature. Gas-phase ΣPCBs (sum of 98 congener groups) are highest in concentration at the Chicago site and lowest at the Sleeping Bear Dunes site. The resulting ΣPCB gas fluxes exhibit a seasonality that reflects elevated summertime gas-phase concentrations not compensated by temperaturecorrected Henry’s law coefficients. Particulate-phase deposition is much smaller in magnitude than gas fluxes, for either compound. Gas and particulate fluxes are comparable only at the Chicago site and only when large (>10 µm) particulates are considered.

Introduction In a continuing effort to understand how chemicals behave in a large lake, the Great Lakes National Program Office (GLNPO) of the U.S. Environmental Protection Agency (U.S. EPA) initiated the Lake Michigan Mass Balance (LMMB) study. The LMMB consists of a large field effort, completed in 1994 and 1995, and an ongoing modeling effort. Both efforts focus on four potentially toxic organic compounds: poly* Corresponding author phone: (319)384-0789; fax: (319)335-5660; e-mail: [email protected]. † The University of Iowa. ‡ State University of New York at Buffalo. 278

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 2, 2001

chlorinated biphenyl congeners (PCBs), trans-nonachlor, atrazine, and mercury. The central goal of the study is to model the behavior of these compounds in Lake Michigan. This requires an estimate of all the loads to and losses from the system and an understanding of the controlling factors for each compound. In previous papers discussing the LMMB data set, we have reported precipitation deposition fluxes of atrazine, nitrogen, and phosphorus (1) and have presented a model for predicting the spatial and temporal variability in gas-phase concentrations of PCBs over the lake (2). PCBs were included in the LMMB, in part, because they continue to contaminate fish in the Great Lakes at a level unfit for human consumption. Efforts to reduce PCB contamination of the Great Lakes and other surface waters include the 1972 ban from production in the United States; the 1978 amendment to the Great Lakes Water Quality Agreement supporting virtual elimination of PCBs from all discharges to the Lakes; and the 1995 amendment to the Clean Water Act (Great Lakes Water Quality Guidance). As a result of the removal of PCBs from industrial processes, waste streams, and other sources, concentrations of PCBs in Lake Michigan water have declined from ∼1.0 ng L-1 in 1980 (3) to ∼0.5 ng L-1 in 1991 (4) to ∼0.1 ng L-1 in 1994 (5). Whether PCBs will continue to decline has been questioned (6) and may depend on a decline of atmospheric loads, a source that has not been adequately quantified. Atmospheric contributions of PCBs to Lake Michigan have been reported for over 20 years (7-10). Since then, investigations using paired air and water data have illustrated that gas exchange is a simultaneous occurrence of volatilization and absorption. Field and modeling studies conducted in the early 1990s found that the lake is either volatilizing PCBs (4, 11, 12) or near chemical equilibrium (13). The Atmospheric Exchange over Lakes and Oceans (AEOLOS) study found that the Chicago-Gary urban industrial region is a major deposition zone for PCBs and other atmospheric pollutants (5, 14-19). Based on these studies, there is large uncertainty and variability (both spatial and temporal) in atmospheric exchange with Lake Michigan, thus establishing the need for a broad sampling program throughout the lake during all seasons. trans-Nonachlor provides an interesting comparison to PCBs. Its physical-chemical properties of water solubility and vapor pressure are very similar to PCBs (20-22), but its sources are quite different. Unlike PCBs, its primary use was agricultural, not industrial, as it was a principal component of the technical mixture of chlordane, an insecticide banned in 1988 (23). Although the insecticide was not intentionally discharged to the lake, its widespread use for commercial, residential, and agricultural purposes may have resulted in its entering the lake through nonpoint sources, including atmospheric exchange. In this paper, we review the PCB and trans-nonachlor data used for estimating the atmospheric loadings for the LMMB modeling effort and present localized results for atmospheric deposition (and volatilization) of PCBs and trans-nonachlor. The intent of the research described in this paper is (a) to assess the quality and trends evident in the atmospheric concentration data, with special attention to anomalies that prevent use of some data for modeling purposes, and (b) to calculate local gas and particulate fluxes of PCBs and trans-nonachlor. Sampling and Analytical Methods. The field effort for the LMMB was conducted over 18 months beginning in April 1994 through October 1995 (23). Over 400 atmospheric samples were collected at eight sampling sites on the shore 10.1021/es991463b CCC: $20.00

 2001 American Chemical Society Published on Web 12/09/2000

FIGURE 1. Shoreline and over-water sampling site locations. around the lake (shoreline) and during seven cruises on the lake (over-water) with the U.S. EPA’s R/V Lake Guardian (Figure 1). These samples were analyzed for persistent organic contaminants. Atmospheric sampling strategies and analytical methods used by the LMMB are detailed in the Lake Michigan Mass Balance Methods Compendium (24, 25) but will be briefly described here. Gas and particulate samples were collected at eight shoreline and 14 over-water sites between April 1994 and October 1995. All samples were analyzed and reported to the U.S. EPA by the Illinois State Water Survey except for those collected at the Sleeping Bear Dunes site, which is operated by the Integrated Atmospheric Deposition Network (IADN). These samples were analyzed by Indiana University after August 1994. A modified high volume (hi-vol) air sampler (Graseby, Cleves, OH) equipped with a 20.3 × 25.4 cm quartz fiber filter (QFF) and XAD-2 resin cartridge was used to collect particulate and gas phases, respectively. During the warm season (May-Oct), air was sampled every 3 days at the Chicago site and every 6 days at the Indiana Dunes, Chiwaukee Prairie, and South Haven sites, and every 12 days at the Beaver Island, Sleeping Bear Dunes, and Manitowoc sites. During the cold season (Nov-Apr), samples were collected every 12 days at the shoreline sites. This resulted in two or three samples per month, which were composited by calendar month and extracted as a whole. Replicate samples were collected using side-by-side samplers and were not composited. Individual filter samples collected during a calendar month were physically combined and a representative subsample was analyzed for the target compounds. At the Sleeping Bear Dunes site, samples were collected every 12 days but were not composited, yielding two or three samples per calendar month. For comparison with composite samples from the other seven shoreline sites, data from individual samples were arithmetically averaged for each month. At each overwater site, a hi-vol was operated continuously for 12 h, while the R/V Lake Guardian held its station. Unlike samples collected at the shoreline sites, gas-phase samples collected at a specific over-water site were not composited. However, particulate samples were composited for sites within close

proximity to each other due to low atmospheric concentrations of total suspended particulates. A higher mass of analyte was required to meet the method detection limit (MDL, Table 1). Samples were extracted, fractionated, and concentrated as described by the Lake Michigan Mass Balance Methods Compendium. There were small differences in the water and air methods that resulted in a different number of congener peaks in the two datasets. Here we describe a list of 98 congener groups that were analyzed by both the water and air methods. The sum of these 98 congeners are described here as ΣPCBs. All sample masses were quantified by the internal standard method and corrected for analytical efficiency using PCB surrogate standards. Quality assurance parameters included surrogate recoveries, matrix spike recoveries, field replicates, and field blanks. Over the 18month sampling period, more than 200 regular and 200 quality assurance gas- and particulate-phase air samples were collected and analyzed for the target compounds (Table 1). With the exception of over-water samples collected in 1994, the quality assurance measures meet or exceed levels of quality expected for the analysis of trace organic compounds in the atmosphere. The over-water samples, which were all collected aboard the R/V Lake Guardian, presented the most important quality assurance challenge to this project. As a result of the analyses described below, all over-water ΣPCB data from 1994 (gasand particulate-phase) were categorized as contaminated and eliminated from subsequent flux modeling. Concentrations of gas-phase ΣPCB and trans-nonachlor measured on the R/V Lake Guardian are illustrated in Figure 2. Aboard the R/V Lake Guardian, over-water air samples were collected at two locations: from a hi-vol sampler secured to the mast on the bow of the ship and from a hi-vol sampler mounted on an extension arm, or yardarm. The yardarm was deployed over the water approximately three meters off the starboard bow. Over the 18 months, at least eight different calibrated hi-vol samplers were used as the motors required repair or replacement. This was discovered during a review of the field notes listing the cumulative operation times recorded for each sampler (see Supporting Information). On 12 occasions, replicate air samples were collected simultaneously. In 1994, the replicate pairs were collected from two samplers secured to the bow mast. In 1995, one sample was collected using the yardarm-mounted hi-vol, the other using the bow-mounted hi-vol. The over-water replicate pairs did not compare well for gas-phase ΣPCBs; the relative percent difference (RPD) was 88.4% (Table 1). The reproducibility was worse than that reported for ΣPCBs measured at the shoreline sites (RPD ) 31.8%) and worse than that reported for gas-phase transnonachlor (RPD ) 38.7% at the shoreline sites and 31.3% at the over-water sites). The mean gas-phase ΣPCB concentration measured in the bow samples was higher (3.0 ng m-3) than measured in the yardarm samples (0.50 ng m-3) and higher than measured at the shoreline sites (0.70 ng m-3 on average). In addition, the samples collected on the yardarm exhibited better agreement with over-water air measurements reported by Zhang et al. (26) (0.65 ng m-3). Their over-water samples were collected on the R/V Lake Guardian using the yardarm-mounted hi-vol sampler. Attempts to identify a source of contamination at the bow have not been successful. Replicate over-water air sampling and surface wipes on the R/V Lake Guardian in 1997 were inconclusive (22). We believe that the bow was contaminated with PCBs during the 1994-1995 sampling season. There is no evidence that the bow was contaminated in earlier years or that samples collected in 1997 or later (yardarm-mounted sampler use) are affected. The problems described above do not apply to gas-phase trans-nonachlor for samples collected in either 1994 or 1995. Over-water concentrations of gasVOL. 35, NO. 2, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

279

TABLE 1. Method Detection Limits (MDL), Surrogate Recoveries, Matrix Spike Recoveries, and Field Replicates for Gas- and Particulate-Phase ΣPCBs and trans-Nonachlora compound

gas

particulate

ΣPCBs trans-nonachlor

MDL 15.6 ng (6.79 × 10-3 ng m-3) -5 0.163 ng (7.08 × 10 ng m-3)

27.9 ng (1.21 × 10-2 ng m-3) 0.454 ng (1.98 × 10-4 ng m-3)

IUPAC No. 14 IUPAC No. 65 IUPAC No. 166

Surrogate Recovery 108.2% ( 20.8% (n ) 196) 88.7% ( 16.0% (n ) 196) 96.8% ( 6.7% (n ) 196)

95.1% ( 9.2% (n ) 163) 81.9% ( 10.7% (n ) 163) 99.1% ( 8.2% (n ) 163)

ΣPCBs trans-nonachlor

Matrix Spike Recovery 96.7% ( 7.0% (n ) 11) 75.2% ( 14.0% (n ) 11)

91.2% ( 7.3% (n ) 19) 72.5% ( 9.2% (n ) 19)

Field Replicates, Relative Percent Difference (RPD) between Two Side-by Side Samplers ΣPCBs shoreline over-water trans-nonachlor shoreline over-water a

31.8% ( 29.0% (n ) 26) 88.4% ( 42.1% (n ) 12)

30.0% ( 33.4% (n ) 20) 32.6% ( 26.4% (n ) 5)

38.7% ( 52.7% (n ) 25) 31.3% ( 48.6% (n ) 12)

68.1% ( 58.9% (n ) 18) 89.3% ( 83.7% (n ) 4)

The MDL concentrations were determined by dividing the MDL mass by the average sample volume (2298 m3).

been reported for semivolatile compounds (27-29). This temporal variation is evident in both in samples collected over the water (Figure 2) and on the shoreline (Figure 3). The variation in ΣPCB concentration is greatest at the Chicago site, where concentrations are about an order of magnitude greater in the summer than in the winter. The variation in trans-nonachlor concentration is largest at the Indiana Dunes site, where the highest concentration is 68 times higher than the lowest concentration. It is likely that the actual ranges are even greater because the sampling methodology was developed to capture mean concentrations, not highs and lows. In a separate paper from our group, Green et al. (2) showed the variation in gas-phase ΣPCB concentrations is controlled by local meteorology and source areas, which caused a predicted variation of 0.17 ng m-3 to 12.6 ng m-3 for ΣPCBs at the Chicago site (a factor of 74).

FIGURE 2. Over-water gas-phase ΣPCB (top) and trans-nonachlor (bottom) concentrations (ng m-3) from samples collected on the R/V Lake Guardian. The southern sites represent sites south of 43° latitude (1, 5, 6, 11, 18M, 380, and 310 in Figure 1). Individual sample site locations and hi-vol identifications, as a function of sample collection order, are provided in the Supporting Information. The x-axis indicates month of collection and is not scaled to time. phase trans-nonachlor are comparable to shoreline concentrations (discussed in the next section). Since all of the samples were collected in 1994 on the R/V Lake Guardian with a bow-mounted sampler, we recommend that the 1994 gas-phase PCB congener data not be used for any atmospheric modeling. Future sampling should be conducted exclusively with the yardarm extensions.

Results and Discussion Gas Concentrations. Gas-phase concentrations of ΣPCBs and trans-nonachlor show a seasonal variation, as has often 280

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 2, 2001

Concentrations of gas-phase ΣPCBs are always higher at the Chicago site (2.5 ng m-3 annual average) than the other sites. The Chicago area is a well-known source of gas-phase and particulate-phase PCBs. No other sites sampled in this study exhibit concentrations as high. For compounds that have historical use in industry, like PCBs, it is important to make measurements near known sources. For example, Milwaukee is a large urban and industrialized area having similar potential sources of anthropogenically derived compounds as Chicago does. There is some evidence of elevated gas-phase PCBs near Milwaukee (12, 30). Unfortunately, samples were not collected in or around Milwaukee during this field study. Beaver Island, one of the most remote sites, had the second highest summertime ΣPCB concentrations (0.79 ng m-3 annual average compared to 0.37 ng m-3 annual average for the remaining six sites). Recent studies of the sampler used at the Beaver Island site indicate that the data was not reflective of over-water concentrations as a result of contamination (31) or unknown site characteristics. For this reason, the Beaver Island data was not considered in the model for atmospheric concentration distribution over the lake (2). Sleeping Bear Dunes, a site which has little seasonal variation and exhibits the lowest concentrations (0.15 ng m-3 annual average concentration), seems to best describe background or upper atmosphere concentrations. Although used by IADN to represent background concentrations, it is not representative of the air over the lake, which has an average annual concentration of 0.46 ng m-3 (2). As discussed below, gas-phase ΣPCB concentrations observed at the Sleeping Bear Dunes site tend to underestimate deposition

FIGURE 4. Particulate-phase ΣPCB (top) and trans-nonachlor (bottom) concentrations (ng m-3) in monthly samples collected between April 1994 and September 1995 at the shoreline sites. The legend is the same as in Figure 3.

FIGURE 3. Gas-phase ΣPCB (top) and trans-nonachlor (bottom) concentrations (ng m-3) in monthly samples collected between April 1994 and September 1995 at the shoreline sites. Discrete data points are connected with a cubic spline to illustrate the temporal trend. The range of concentrations for over-water samples collected on the R/V Lake Guardian (labeled OW94 and OW95) are shown as box plots. Replicate samples are not included in this figure. fluxes to Lake Michigan and overestimate volatilization fluxes when used to estimate exchange with the entire lake. Gas-phase trans-nonachlor concentrations decrease with increasing latitude. Gas-phase concentrations are higher at the southern shoreline sites (16 pg m-3 annual average) of Muskegon, South Haven, Indiana Dunes, Chicago, and Chiwaukee Prairie than at the northern shoreline sites (5.7 pg m-3 annual average) of Beaver Island, Sleeping Bear Dunes, and Manitowoc. The variation in concentrations is statistically correlated with latitude and temperature. These parameters were examined using a multiple linear regression of all the gas-phase trans-nonachlor concentrations (ln Cg ) 43-9358 T-1 - 0.198 latitude; R 2 ) 0.73, p < 0.0001 for all coefficients). The remaining variability may be due, in part, to atmospheric dilution over the lake and spring tillage. Gas-phase concentrations in samples collected aboard the R/V Lake Guardian are usually lower than those measured in samples collected at the shoreline sites. A figure supplied in the Supporting Information illustrates the interpolated concentrations of gas-phase trans-nonachlor in May 1995. During this month, the southern shoreline sites exhibited high concentrations of gas-phase trans-nonachlor that cannot be explained by temperature and was not observed for the PCB

congeners. We suggest that this is a signal from spring tillage but cannot confirm this hypothesis. Particulate Concentrations. Particulate-phase ΣPCBs and trans-nonachlor concentrations measured during the 19941995 field season demonstrate less predictable temporal variability than gas-phase concentrations (Figure 4). However, like gas-phase ΣPCBs, particulate-phase ΣPCB concentrations from samples collected at Chicago are consistently higher than any of the other sites (Table 2). This is partly a result of the higher concentration of atmospheric particulates in Chicago, which contribute to the higher volume-normalized concentrations measured there. Total suspended particulate (TSP) concentrations averaged 66 µg m-3 annually in Chicago and averaged 34 µg m-3 annually at the other seven shoreline sites. This is somewhat higher than reported by Franz et al. (16), who sampled during the same time of year as this study and found TSP concentrations to be 27 µg m-3 in Chicago and 15 µg m-3 over the lake. Gas-particle distributions were not evaluated for the LMMB dataset because of the difficulty in matching gas- and particulate-phase data, which were separately composited over several days, with the appropriate TSP. A more complete consideration of gas-particle distributions of persistent organic pollutants in the Chicago area can be found in Simcik et al. (15). Local Gas Fluxes. The modified two-film, gas exchange model has been described elsewhere (11, 12, 22, 32) and is described briefly here. The net flux (absorption less volatilization) of a chemical across the air-water interface is calculated using a mass transfer coefficient and a concentration gradient between the air and the water (33)

(

Fg ) kol Cw -

)

CgRT H

(1)

where Fg is the net gas exchange flux (ng m-2 mo-1); kol (m VOL. 35, NO. 2, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

281

TABLE 2. Mean Atmospheric Concentrations and Mean Instantaneous Fluxes of ΣPCB and trans-Nonaclor in Lake Michiganc over-water (R/V Lake Guardian)

n

shoreline excluding Chicago

n

Chicago

2600 ( 1900 -14000 ( 13000

n

Gas-Phase Concentrations and Net Exchange ΣPCBs Cg (pg m-3) Fg (ng m-2 mo-1) trans-nonachlor Cg (pg m-3) Fg (ng m-2 mo-1)

970 ( 1100 -300 ( 4200

42 35

420 ( 440 -320 ( 2800

126 126

7.3 ( 10 46 ( 67

69 69

11 ( 12 42 ( 31

126 126

21 ( 19 13 ( 35

18 18 18 18

Particulate-Phase Concentrations and Deposition ΣPCBs Cpt (pg m-3) Fpt (ng m-2 mo-1)

20 ( 15 100 ( 76

14 14

24 ( 19 120 ( 98

126 126

85 ( 49 440a ( 250 16000b ( 900

18 18 see text

trans-nonachlor Cpt (pg m-3) Fpt (ng m-2 mo-1)

0.31 ( 0.35 1.6 ( 1.8

25 25

0.45 ( 0.37 2.3 ( 1.9

124 124

1.2 ( 0.9 6.2a ( 4.7 230b ( 180

18 18 see text

a Calculations without consideration of large particles. b Calculations with consideration of large particles. c Net gas deposition fluxes are indicated by a negative (-) sign. Standard deviations are indicated by (.

mo-1) is the overall mass transfer coefficient; Cw (ng m-3) is the dissolved-phase concentration; Cg (ng m-3) is the gasphase concentration; R (atm m3 mol-1 K-1) is the universal gas constant; T (K) is the interfacial temperature; and H (atm m3 mol-1) is the Henry’s law coefficient corrected for the interfacial temperature. The mass transfer coefficient was estimated using empirical correlations from field and laboratory experiments (34-41) and determined for the mean wind speed measured at the sampling site during the sampling period. The Henry’s law coefficient has a large effect on the magnitude and direction of gas exchange calculation. For trans-nonachlor, H at 25 °C is 3.1 × 10-4 atm m3 mol-1 and was corrected for surface water temperature using an enthalpy value of 63.3 kJ mol-1 (21). For individual PCB congeners, gas exchange fluxes were calculated using two different sets of H values. Fluxes were first estimated using the empirical relationship derived by Brunner et al. (42) for each of the 98 individual congeners (Figure 5) and corrected for temperature using an enthalpy of 50 kJ mol-1 (43). Fluxes were then estimated using the H and enthalpy values measured by Bamford et al. (44) for 20 of the 98 congeners combined with values from the Brunner et al. relationship for the remaining congeners. Gas exchange fluxes for only ΣPCBs (sum of 98 individual congeners fluxes) are reported here. The substitution of these 20 H values do not change the overall conclusions for ΣPCB fluxes, although the Bamford et al. values do predict more gross volatilization when examined for any specific congener. Congener-specific PCB and trans-nonachlor dissolved water concentrations were taken from the water sampling site closest to the corresponding shoreline site (see table in Supporting Information). The air samples collected overwater were paired with the water samples collected at the same site. There was less temporal variability in dissolved water ΣPCB (0.17 ( 0.09 ng L-1) than in trans-nonachor (5.5 ( 12 pg L-1) concentrations (22). The over-water fluxes (Figure 6) were calculated using concentrations from the spatially and temporally paired air and water samples. These fluxes are therefore “instantaneous” and specific to one day and one location on the lake. The fluxes calculated for the shoreline sites (Figure 5) used an average of all dissolvedphase water concentrations at the closest site. Gas exchange fluxes were calculated for each PCB congener, summed, and reported as ΣPCB fluxes for each of the eight shoreline (Figure 5) and 14 over-water sites (Figures 6). trans-Nonachlor gas exchange fluxes were calculated 282

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 2, 2001

similarly, although only the over-water sites are illustrated (Figure 6). The over-water ΣPCB flux predictions may be compared directly to Zhang et al. (26) and to Hornbuckle et al. (12) and Hillary et al. (13) who calculated gas exchange fluxes at two of the sites investigated here. These previous reports for the Chicago region and for the Sleeping Bear Dunes site provide a good comparison to this work because they represent very different site characteristics. Zhang et al. report over-water gas exchange fluxes for the Chicago area in 1994-1995 ranging from -32 (deposition) to +46 (volatilization) ng m-2 day-1. At comparable near-Chicago sites (1 and 5), we report a range of -310 to + 85 ng m-2 day-1 (Figure 6). This difference in range is due, in part, to the number of congeners represented as ΣPCBs (98 congeners in this study versus 25 in Zhang et al.). We also measured several events of high over-water gas-phase concentrations (Cg > 2 ng m-3). These events are shown as outliers in the box plots in Figure 3. When adjusted for number of congeners and the Cg outliers are excluded, our mass transfer coefficients, water concentrations, air concentrations, and resulting fluxes agree with Zhang et al. other investigators have reported ΣPCB gas exchange fluxes for the Sleeping Bear Dunes site from samples collected in 1991-1992 and in 1993-1994 (12, 13). Both the water- and gas-phase concentrations are lower in 19941995 than in 1991-1992 (p value 3 standard deviations) than the whole lake mean ΣPCB concentration. Only fluxes calculated with Henry’s law coefficients from Brunner et al. (42) are plotted. The x-axis is the same as in Figure 2. Site identification is listed in the Supporting Information. urban-industrial sources is difficult to estimate, despite extensive fieldwork completed for this study, the Lake Michigan Urban Air Toxics (LMUAT) in 1991 and AEOLOS in 1994-1995. The area that the large particles can be deposited is highly variable and depends on mixing height, wind speed and direction, and particle concentration. In a separate paper, we reported that PCBs originating in Chicago influence an average annual area bounded within a 40 km radius that does not extend to he southern tip of the lake (2). Assuming that particulate-phase ΣPCBs are influenced by factors similar to gas-phase ΣPCBs and also considering the much larger removal efficiencies for large particles, we estimate that the average area of impact can be no more than 100 km2 surrounding the Chicago site. The resulting whole lake annual flux is then 100 kg yr-1 for ΣPCBs, 20 kg yr-1 due to the Chicago site alone. For trans-nonachlor, the whole lake annual load is approximately 2 kg yr-1, 0.3 kg yr-1 due to the Chicago site. This is three times larger than estimated by Hillary et al. (13), who estimated fluxes using background concentrations from samples collected at the Sleeping Bear Dunes site. Although we believe our value is reasonable, and based on a very large sample set, the difficulties with interpretation of the data for large particles are significant. Implications. One of the goals of the LMMB is to guide decisions about reducing inputs of persistent organic compounds to Lake Michigan. As a result of the work described here and elsewhere, it is clear that urban/industrial areas are large sources of atmospheric PCBs. The magnitude of the impact on the whole lake requires an atmospheric model that is not applied here but is an ongoing component of the project. Reduction strategies should focus on the identification of major volatilization sources in the Chicago region 284

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 2, 2001

The concentrations of PCBs and trans-nonachlor presented here have been approved and released by the U.S. Environmental Protection Agency, Great Lakes National Program Office. This research was supported by a grant from the U.S. Environmental Protection Agency, Great Lakes National Program Office (GL985872-01-0), Project Officer Angela Bandemehr. This paper benefited from discussions with Clyde Sweet at the Illinois State Water Survey, who as well as Karen Harlin at the Illinois State Water Survey and Ilora Basu and Ron Hites at Indiana University provided preliminary chemical data prior to the EPA release. We also thank Marcia Kuehl, contractor for the U.S. EPA for assistance with interpreting the format used in the data set.

Supporting Information Available Tables of gas-phase shoreline, gas-phase over-water, particulate-phase shoreline, and particulate-phase over-water and a plot of trans-nonachlor in air over Lake Michigan. This material is available free of charge via the Internet at http:// pubs.acs.org.

Literature Cited (1) Miller, S. M.; Sweet, C. W.; DePinto, J. V.; Hornbuckle, K. C. Environ. Sci. Technol. 2000, 34, 55-61. (2) Green, M. L.; Depinto, J. V.; Sweet, C.; Hornbuckle, K. C. Environ. Sci. Technol. 2000, 34, 1833-1841. (3) Swackhamer, D. L.; Armstrong, D. E. J. Great Lakes Res. 1987, 13, 24-36. (4) Pearson, R. F.; Hornbuckle, K. C.; Eisenreich, S. J.; Swackhamer, D. L. Environ. Sci. Technol. 1996, 30, 1429-1436. (5) Offenberg, J. H.; Baker, J. E. Environ. Sci. Technol. 1997, 31, 1534-1538. (6) Smith, D. W. Environ. Sci. Technol. 1995, 29, 42A-46A. (7) Murphy, T. J.; Rzeszutko, C. P. J. Great Lakes Res. 1977, 3, 305312. (8) Doskey, P. V.; Andren, A. W. J. Great Lakes Res. 1981, 7, 15-20. (9) Murphy, T. J.; Pokojowczyk, J. C.; Mullin, M. D. In Physical Behavior of PCBs in the Great Lakes; Mackay, D., Paterson, S., Eisenreich, S. J., Simmons, M. S., Eds.; Ann Arbor Science: Ann Arbor, MI, 1983; pp 49-58. (10) Swackhamer, D. L.; Armstrong, D. E. Environ. Sci. Technol. 1986, 20, 879-883. (11) Achman, D. R.; Hornbuckle, K. C.; Eisenreich, S. J. Environ. Sci. Technol. 1993, 27, 75-87. (12) Hornbuckle, K. C.; Sweet, C. W.; Pearson, R. F. Environ. Sci. Technol. 1995, 29, 869-77.

(13) Hillery, B. R.; Simcik, M. F.; Basu, I.; Hoff, R. M.; Strachan, W. M. J.; Burniston, D.; Chan, C. H.; Brice, K. A.; Sweet, C. W.; Hites, R. A. Environ. Sci. Technol. 1998, 32, 2216-2221. (14) Simcik, M. F.; Zhang, H. X.; Eisenreich, S. J.; Franz, T. P. Environ. Sci. Technol. 1997, 31, 2141-2147. (15) Simcik, M. F.; Franz, T. P.; Zhang, H. X.; Eisenreich, S. J. Environ. Sci. Technol. 1998, 32, 251-257. (16) Franz, T. P.; Eisenreich, S. J.; Holsen, T. M. Environ. Sci. Technol. 1998, 32, 3681-3688. (17) Paode, R. D.; Sofuoglu, S. C.; Sivadechathep, J.; Noll, K. E.; Holsen, T. M.; Keeler, G. J. Environ. Sci. Technol. 1998, 32, 1629-1635. (18) Zufall, M. J.; Dai, W. P.; Davidson, C. I. Atmos. Environ. 1999, 33, 4283-4290. (19) Zhang, H. X.; Eisenreich, S. J.; Franz, T. R.; Baker, J. E.; Offenberg, J. H. Environ. Sci. Technol. 1999, 33, 2129-2137. (20) Mackay, D.; Shiu, W. Y.; Ma, K. C. Illustrated Handbook of Physical-Chemical Properties and Environmental Fate for Organic Chemicals; Lewis Publishing Co.: Chelsea, MI, 1992; Vol. 1: Monoaromatic hydrocarbons, chlorobenzenes and PCBs. (21) Iwata, H.; Tanabe, S.; Ueda, K. Environ. Sci. Technol. 1995, 29, 792-801. (22) Miller, S. M. M.S. Thesis, State University of New York at Buffalo, 1999. (23) Horvatin, P. J. Lake Michigan mass balance budget/mass balance work plan; U.S. Environmental Protection Agency/Great Lakes National Program Office: October 14, 1993. (24) USEPA. Lake Michigan Mass Balance Study (LMMB) Methods Compendium, Volume 1: Sample Collection Techniques; EPA905/R-97-012a; U.S. EPA, Great Lakes National Program Office: Chicago, IL. (25) USEPA. Lake Michigan Mass Balance Study (LMMB) Methods Compendium, Volume 2: Organic and Mercury Sample Analysis Techiques; EPA-905/R-97-012b; USEPA, Great Lakes National Program Office: Chicago, IL. (26) Zhang, H.; Eisenreich, S. J.; Franz, T. R.; Baker, J. E.; Offenberg, J. H. Environ. Sci. Technol. 1999, 33, 2129-2137. (27) Hoff, R. M.; Muir, D. C. G.; Grift, N. P. Environ. Sci. Technol. 1992a, 26, 266-275. (28) Burgoyne, T.; Hites, R. Environ. Sci. Technol. 1993, 27, 910914. (29) Cortes, D. R.; Basu, I.; Sweet, C. W.; Brice, K. A.; Hoff, R. M.; Hites, R. A. Environ. Sci. Technol. 1998, 32, 1920-1927. (30) Bogdan, J. J. M.S. Thesis, State University of New York at Buffalo, 1999. (31) Basu, I.; O’Dell, J.; Arnold, K.; Hites, R. Environ. Sci. Technol. 2000, 34, 356-360.

(32) Hornbuckle, K. C.; Sweet, C. W.; Eisenreich, S. J. Environ. Sci. Technol. 1994, 28, 1491-501. (33) Liss, P.; Slater, P. Nature 1974, 247, 181-184. (34) Schwarzenbach, R. P.; Gschwend, P. M.; Imboden, D. M. WileyInterscience: New York, 1993; p 199. (35) Fuller, E. N.; Schettler, P. D.; Giddings, J. C. Ind. Eng. Chem. 1966, 58, 19-27. (36) Mackay, D.; Yuen, A. T. K. Environ. Sci. Technol. 1983, 17, 211217. (37) Upstill-Goddard, R.; Watson, A.; Liss, P.; Liddicoat, M. Tellus 1990, 42B, 364-377. (38) Wanninkhof, R.; Ledwell, J. R.; Broecker, W. S. Science 1985, 227, 1224-1226. (39) Wanninkhof, R. H.; Asher, W.; Weppernig, R.; Chen, H.; Schlosser, P.; Langdon, C.; Sambrotto, R. J. Geophys. Res. 1993, 98, 2023720248. (40) Wanninkhof, R.; Ledwell, J. R.; Crusius, J. In Air-Water Mass Transfer; Wilhelm, S. C., Gulliver, J. S., Eds.; American Society of Civil Engineers: New York, 1991; pp 441-458. (41) Wilke, C. R.; Chang, P. A. I. Ch. E. Journal 1955, 1, 264-270. (42) Brunner, S.; Hornung, E.; Santl, H.; Wolff, E.; Piringer, O. G.; Altschuh, J.; Bruggemann, R. Environ. Sci. Technol. 1990, 24, 1751-1754. (43) Hulscher, T. E. M.t.; Velde, L. E.v.d.; Bruggeman, W. A. Environ. Toxicol. Chem. 1992, 11, 1595-1603. (44) Bamford, H. A.; Poster, D. L.; Baker, J. E. J. Chem. Eng. Data in press. (45) Holsen, T. M.; Noll, K. E.; Fang, G.-C.; Lee, W.-J.; Lin, J.-M. Environ. Sci. Technol. 1993, 27, 1327-1333. (46) Williams, R. M. Atmos. Environ. 1982, 16, 1933-1938. (47) Caffrey, P. F.; Ondov, J. M.; Zufall, M. J.; Davidson, C. I. Environ. Sci. Technol. 1998, 32, 1615-1622. (48) Offenberg, J. H. Ph.D. Dissertation, University of Maryland, 1998. (49) Landis, M. S. Ph.D. Dissertation, University of Michigan, 1998. (50) Eisenreich, S. J.; Strachan, W. M. J. Estimating Atmospheric Deposition of Toxic Substances to the Great Lakes, an Update; Environment Canada, Canada Centre for Inland Waters: Burlington, Ontario, January 31-February 2, 1992; p 60.

Received for review December 31, 1999. Revised manuscript received September 26, 2000. Accepted October 12, 2000. ES991463B

VOL. 35, NO. 2, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

285