Volatilization of Polychlorinated Biphenyls from Green Bay, Lake Michigian Dlane R. Achman,? Kerl C. Hornbuckle,+and Steven J. Eisenreich",'
Gray Freshwater Biological Institute and Department of Civil and Mineral Engineering, University of Minnesota, P.O. Box 100, Navarre, Minnesota 55392
rn The volatilization of polychlorinated biphenyls (PCBs)
from Green Bay was estimated as part of the Green Bay Mass Balance Study (U.S.EPA). The strategy employed was to simultaneously collect air and water samples above and below the air-water interface, analyze the atmospheric gas phase and the water column dissolved phase for 85 PCB congeners, and calculate the direction and magnitude of flux for each congener using Henry's law and meterological and hydrological parameters. Sampling covered the period of June through October 1989. Air-water transfer rates were calculated for the 14 individual days spanning the three cruises by using the stagnant two-film model. Calculated total PCB volatilization rates ranged from 13 to 1300 ng/m2.day. The most important factors affecting the magnitude of the flux are wind speed and water concentration. The range of fluxes calculated compares well with other estimates for the Grate Lakes. The results of this study support the hypothesis that volatilization is an important phenomenon controlling the fate of hydrophobic organic chemicals (HOCs) in aquatic systems.
Introduction The Great Lakes ecosystem has suffered severely as a result of anthropogenic inputs of a wide variety of toxic organic chemicals. Polychlorinated biphenyls (PCBs) were discovered in the water, sediment, and fish of Green Bay, Lake Michigan, in the early 1970s (I). Green Bay is a long, relatively shallow extension of Lake Michigan with a maximum depth of 30-35 m. It has been characterized as an estuary by Smith et al. (I)because of its high biological activity and chemical and thermal differences from Lake Michigan, although marine scientists dispute this characterization. The mixing processes in the bay are also similar to those of an estuary. The southern one-third of the bay varies greatly from the northern two-thirds in terms of ita chemical and biological characteristics. The southern portion is hypereutrophic and the northern portion is mesotrophic to oligotrophic, similar to Lake Michigan. Sources of PCBs to an aquatic ecosystem may include tributary inputs, municipal and industrial discharges, and atmospheric deposition. Major losses include sedimentation,volatilization, and outflow. The importance of each input process and loss process varies among each of the Great Lakes. The most important source of PCBs to Green Bay is the Fox River (1-3). Thirteen paper mills and five major municipal wastewater treatment facilities Department of Civil and Mineral Engineering. Gray Freshwater Biological Institute. 0013-936X/93/0927-0075$04.00/0
line the banks of the Fox River ( I ) . Swackhamer and Armstrong (2)reported the highest PCB concentration, 121 ng/L, in the lower bay and concentrations decreasing to 1-2 ng/L with increasing distance northward from the mouth of the Fox River. Contaminated sediments from the Fox River that have been historically deposited in the bay are also thought to contribute a significant amount of PCBs annually (I,3). PCB losses due to volatilization may be of the same magnitude as losses to the sediment for the Great Lakes (4-6). It is intuitive that volatilization is an important phenomenon controlling the fate of many low-boiling (i.e., semivolatile)compounds, but volatilization may also be important for high-boiling chemicals with low aqueous solubilities and vapor pressures such as PCBs (7-22). Mass balance approaches have been employed to determine the overall importance of volatilization as a loss function of PCBs to the Great Lakes (2,4,5,13,14). The objective of this project was to quantify the direction and magnitude of air-water exchange of PCBs in Green Bay, Lake Michigan. The overd strategy was to simultaneously collect air and surface water samples from the bay, analyze them for polychlorobiphenyls, and determine a concentration gradient so that a flux across the air-water interface could be determined.
Experimental Section Air measurements included collection of gas-phase, aerosol, and total suspended particulate (TSP) samples. Water measurements included filtered and dissolved-phase PCBs as well as dissolved organic carbon (DOC), total suspended matter (TSM), and particulate organic carbon (POC) samples. Samples were collected from the R/V Bluewater on the Green Bay Mass Balance Study (GBMBS) cruises in the summer and fall of 1989 during the periods of June 4-14, July 28-August 1, and October 20-23. Samples were taken at either four or five sites on each cruise. The sampling sites were numbered 4, 10, 14, 18, and 21 and spanned the bay from south to north (Figure 1). Site 4 is the most southerly site and is closest to the mouth of the Fox River and the city of Green Bay. Table I shows physical characteristics of each site. Marti and Armstrong (3)found that the Fox River is the major source of PCBs to the bay. Smith et al. (I)reported useful information about the history of contamination of Green Bay. Surface water was sampled from the R/V Bluewater by pumping water from a 1-2-m depth through 25.4-mm-0.d. Tygon tubing directly through a precleaned 293-mm glass fiber filter (Schleicher and Schuell No. 25,0.7-pm nominal
0 1992 American Chemical Society
Environ. Sci. Technol., Vol. 27, No. 1, 1993
75
GREEN BAY
M I C H I G A N
U
W I S C O N S I N
M I C H I G A N
Figure 1. Water and air sampling locations, Green Bay, 1989.
Table I. Green Bay Sampling Site Characterization
site 4 10 14 18 21
water temp at 2 m (‘‘2) June July-Aug Oct
latitude/ longitude
depth (m)
44’36’22” 087’57’16” 44’42’37” 087’51’19” 44’51’02” 087’43’40” 45’05’27” 087’24’41’’ 45O17’36” 087O09”54”
6
11.6-12.1
15.3
6.9
9
9.0-10.1
15.6
7.6
15
9.2
15.2
8.3
30
6.0-7.1
14.4
6.7
27
14.0
pore size) housed in a stainless steel filter head (293 mm). Prior to use, the glass fiber filters were combusted at 450 “C overnight and transported to the field in aluminum foil. The flow rate through the filter was held at -10 L/min, and 40-200 L of water were filtered. During the middle of the filtration, 65 L of the filtrate was collected in a precleaned stainless steel holding tank and then passed through a 2.5 cm i.d. X 20 cm column of Amberlite XAD-2 resin (Sigma Chemical Co., A-7643) at 250 mL/min to isolate dissolved-phase PCBs from the water (5,15).In some cases the surrogate standards were added to the resin column prior to analyta isolation to examine the recovery efficiency of XAD-2 from this stage of the procedure. The resin was precleaned by sequential 24-h extractions with methanol, acetone, hexane, and dichloromethane, followed by 24-h extractions with hexane, acetone, and methanol to ensure a wettable surface. The resin was rinsed with Milli-Q water and stored in amber solvent bottles filled with Milli-Q water. The water was repeatedly decanted to remove the methanol. Columns packed with clean XAD-2 resin that were transported to the field and returned to the laboratory unopened served as field blanks. Filter samples were wrapped in aluminum foil, placed in a plastic bag, and frozen until analysis. XAD-2 samples were transferred from the columns to glass jars in the 78 Envlron. Scl. Technol., Vol. 27, No. 1, 1993
laboratory and refrigerated until analysis. The operationally defined particulate phase (PCBs isolated by filtration) and dissolved phase (PCBs isolated using XAD-2 resin) in water were extracted using the same procedure. Samples were placed in a Soxhlet apparatus and rinsed with 200-500 mL of methanol to remove excess water associated with the samples. The methanol rinse was saved and replaced with clean methanol. Surrogate standards containing 3,5-dichlorobiphenyl(IUPAC no. 14), 2,3,5,6-tetrachlorobiphenyl (IUPAC no. 65), and 2,3,4,4’,5,6-hexachlorobiphenyl(IUPAC no. 166) were added a t this point if they were not previously added in the field. The samples were extracted for 24 h each with methanol and dichloromethane. The three resulting fractions (methanol rinse, methanol, and dichloromethane) were combined, and the methanol was removed by liquid-liquid partitioning into water. The combined dichloromethane extract was reduced to -4 mL and the solvent switched to hexane using a Kuderna-Danish apparatus and further reduced to 2 mL under a prepurified nitrogen stream. The sample was then split into two fractions, 25 % for polycyclic aromatic hydrocarbon analysis and 75% for PCB analysis. The remaining PCB fraction was eluted from a column containing 3 g of 6% water-deactivated silica and 10 g of 10% water-deactivated alumina with 60 mL of hexane. The hexane fraction was once again reduced to 4 mL using a Kuderna-Danish apparatus and further reduced to 2 mL under a prepurified nitrogen stream. Samples were then transferred to 4-mL amber vials and internal standards 2,4,64richlorobiphenyl (IUPAC no. 30) and 2,2’,3,4,4’,5,6,6’-octachlorobiphenyl (IUPAC no. 204) were added prior to final solvent reduction by nitrogen blowdown to 100-pL volume. The average percent recoveries for the entire analytical procedure for PCB congener no. 14, no. 65, and no. 166 were 81.0 f 21.8, 96.4 f 17.0, and 95.1 f 15.2, respectively, for all water samples (dissolved and particulate) analyzed in this study ( N = 61). Air samples were collected concurrently with water samples using a General Metal Works high-volume air sampler mounted on the bow of the R/V Bluewater at -3 m above the water surface. The vessel remained anchored at the sampling site until the air and water sampling was complete, usually 7-10 h. Air (300-500 m3) was passed through a glass fiber filter (Schleicher and Schuell no. 25, 20 X 25 cm) and polyurethane foam (PUF) plug (8 cm diameter X 10 cm height) in series at a rate of 0.5-0.7 m3/min. Ancillary measurements including temperature, wind speed, and surface wind direction were collected through the sampling period. Details of sampling and analytical procedures can be found in Hornbuckle et al. (16). PCB samples were analyzed on a Hewlett-Packard 5890A gas chromatograph equipped with an electron capture detector (ECD) and a 30 m X 0.32 mm i.d. dimethyl diphenyl polysiloxane capillary column with a 0.25-~mfilm thickness (DB-5, J&W Scientific). Analysis conditions were as follows: spitless injection (2 pL) 18-kPa helium carrier gas flow; temperature program, injection at 100 “C, held for 10 min, then temperature increased 10 “C/min to 160 “C, then 0.5 “C/min to 227 “C followed by 10 OC/min to 260 “C. The injector temperature was 250 “C, and the ECD was held at 300 “C. Chromatographic data were collected and integrated using the Dynamic solutions Maxima 820D data system (Waters-Millipore). A mixed Aroclor calibration standard containing 250 ng/mL 1232,180 ng/mL 1248, and 180 ng/mL 1262 was prepared to yield a PCB concentration of 610 ng/mL.
-
Congener masses were calculated from the known total PCB concentration of the calibration standard and congener composition of the standard as listed in Mullin et al. (17). Response factors relative to the internal standard (RRF) were generated each day. Individual congener masses were calculated using the RRF and internal standard remonse in the samde according- to GBMBS guidelines (1s). In order to make data comparable within the GBMBS, samples were corrected for surrogate recovery. The percent recovery of surrogate standard PCB no. 14 was used to correct the concentrations of di- and trichlorinated biphenyls, the percent recovery of surrogate standard PCB no. 65 was used to correct the concentrations of tetra- and pentachlorinated biphenyls, and the percent recovery of surrogate standard PCB no. 166 was used to correct the concentrations of hexa- through nonachlorinated biphenyls
dissolved PCB concentration in the bulk water, and C* ( P / H , mol/m3) is the air concentration expressed as a water concentration in equilibrium with the air. The variable P is the vapor-phase air concentration measured (m0l/m3)and converted to units of pressure using the ideal gas law, and H is Henry’s law constant (atm.m3/mol). The with units of velocity overall mass-transfer coefficient is KO] (m/day). The concentration gradient determilies the direction of flux and drives the mass transfer whereas Kol is a kinetic parameter which quantifies the rate of transfer. The value of KO, is dependent on the physical and chemical properties of the compound as well as environmental conditions. The reciprocal of Kolis the total resistance to transfer expressed on a gas- (RT/Hk,) and liquid- (l/k,) phase basis: l/KoI = l / k w + RT/Hk, (2)
(19).
where k, is the water-side mass-transfer coefficient (m/ day) and k, is the air-side mass-transfer coefficient (m/ day). H is Henry’s law constant, R is the universal gas atm.m3/mol K), and T is the abconstant (8.2057 X solute temperature, K. There have been two main approaches to obtaining transfer velocities: laboratory-scale wind tunnel experiments and field measurements. Most experiments, whether laboratory or field studies, are aimed at describing the relationship between the magnitude of gas transfer and wind speed. The tracers used in these experimenta are liquid-phase-controlle&therefore the investigators assume k, is negligible and k, is dominant. The most extensive laboratory studies have employed wind-wave tanks and measured the flux of a variety of gases across the air-water interface such as C02,O2 (26),and organic compounds (27). Field studies include the measurement of the flux of natural or artificial tracer gas across the interface. Until recently, most measurements of gas exchange in lakes were based on the isotopes 14C,222Rn,and 3He (e.g., refs 28 and 29). More recently sulfur hexafluoride (SF,) has been added to lakes and used as a tracer of gas transfer because it has a low aqueous solubility, is inert, and is detectable at low concentrations (30-33). Some investigators now employ a dual-tracer technique where the change in ratio of the tracer concentrations is measured with time (33). Liss and Merlivat (24) summarized the transfer coefficient as a function of wind speed reflecting the information provided by wind tunnel experiments. In the range of environmental wind speeds (0-17 m/s) there are three regimes where different physical processes appear to be controlling gas exchange. The smooth surface regime (wave-free) covers wind speeds of 5 f 3 m/s with the value of k, increasing only very gradually. At wind speeds of 5-10 m/s, capillary waves are formed and the transfer velocity increases considerably. Above wind speeds of 10 m/s, gas transfer is further increased by bubbles formed by breaking waves. Liss and Merlivat’s (24) relationships based on laboratory and field data (30) for the variation of k, with wind speed were used in our calculations of KO$
Field blanks and procedural blanks were analyzed to determine possible sources of contamination. Field blanks consisted of a clean sampling matrix (XAD-2 column, PUF plug, etc.) transported to the field and returned to the laboratory unopened. Field blanks exhibited slightly higher concentrations of target compounds than laboratory blanks. Filter blanks were virtually free of contamination. Procedural blanks (solvent containing surrogate standards) extracted and carried through the analytical procedure were generally free of PCB congener contamination except for coeluting contaminants at PCB IUPAC congeners no. 1, no. 3, and no. 198. These peaks were subtracted from all samples. Compared to total masses of PCBs in water samples, the masses of analytes in the blanks were insignificant. Air sample blanks (solvent, PUF, or XAD-2) contained additional contaminants which coelute with PCB congeners; therefore, six more congeners were eliminated in air samples. Theory and Calculations The model most frequently used to define the transfer of chemicals across the air-water interface is the two-resistance of film model developed by Whitman (20)and popularized by Liss and Slater (21). According to this model, the air-water interface consists of stagnant gas and liquid films adjacent to the interface. The bulk water and air compartments are assumed to be well-mixed. The rate of transfer is governed by molecular diffusion through the stagnant boundary layer. The two-film model was used in this study and will be only briefly expanded upon here because the model has been extensively reviewed (21-24). The thickness of the films vary spatially and temporally (21). Depending on the compound, liquid-phw resistance, gas-phase resistance, or some combination of the two determines the rate of transfer. Resistance to mass transfer occurs primarily in the liquid phase for compounds with atm-m3/molwhile Henry’s law constants (H)of >4.4 X the gas phase controls transfer of hydrophobic organic atm-m3/mol (11). chemicals (HOCs) with H < 1.2 X The bulk water and air compartments are assumed to be well-mixed, and transfer of the compound to the interface through the bulk fluid is not rate-limiting because turbulent transport exceeds molecular transport by several orders of magnitude in the bulk phase (25).The rate of PCB exchange is limited mainly by the liquid film, but the gas film also exhibits some resistance to transfer. The basic equation used to describe the rate of transfer across the interface is F K,I(C, - C*) (1) where F is the flux (mol/m2-day), C, (mol/m3) is the
-
k, = 0 . 1 7 ~for ~ ~ul0 < 3.6 m/s k, = 2 . 8 5 -~ 9.65 ~ ~ for 3.6 < ul0 < 13 m/s
(4)
k, = 5.9u10- 49.3 for ul0 > 13 m/s
(5)
(3)
where ul0 is wind speed at a reference height of 10 m (in units of m/s) and k, has units of centimeters per hour. These relationships have been normalized to C02,which has a Schmidt number (Sc) of 600 at 20 “C (Schmidt number is the kinematic viscosity of the medium divided by the compound’s diffusivity through the medium). Envlron. Scl. Technol., Vol. 27, No. 1, 1993 77
-
Table XI. Calculated Diffusivities and Schmidt Numbers of PCBs in Water and Air for a Dissolved Water Sample from Site 4
r a n g e of wind speeds
degree of chlorination
observed in Green Bay
0
di tri tetra penta hexa hepta octa
I
I
2
4
6
8
10
12
14
16
Wind speed m / s
Figure 2. Liss and Merllvat’s (24)relationship for the variation of k , with wind speed (at a height of 10 m) for a gas with a Schmidt number of -600. The three regions wRh different slopes correspond to the three equations discussed in the text.
Figure 2 shows the relationships listed above and the range of wind speeds observed in Green Bay. Field experiments by Watson et al. (33) using a dual-tracer technique to measure the transfer of dissolved gas across the air-sea interface support the above wind speed dependence suggested by Liss and Merlivat (24). A correction is needed to adjust the above equations if the Schmidt number for the compound is different from 600: kw(COz)/kw(PCB) = SCn(COz)/SCn(PCB) (6) where
n = -y3 for utO< 3.6 m/s and n = -f/z for ul0 > 3.6 m/s The Schmidt number for PCBs was determined by first calculating the diffusivity of PCBs through water using Wilke and Chang’s method (34) and then dividing the kinematic viscosity of water (cm2/s) by the diffusivity of PCBs through water. The Schmidt number for PCB congeners ranges from 2000 to 5000 for di- through octachlorobiphenyls. Equation 3 suggests that during a nowind condition k , would be zero. This probably is not true because, at the very least, molecular diffusion controls gas transfer through the stagnant water film. Diffusion through the water film can be calculated, but the thickness of the film must be known. Schwarzenbach et al. (12) suggested a quadratic relation for no-wind conditions which yields a k , of 0.34 m/day and implies a stagnant water film of about 20-200-pm thickness. Mackay and Yuen (27) gave a “still air” value for k , of 0.053 m/day, which corresponds to a film thickness of 3.2 mm. Using the diffusion of a pentachlorinated PCB in water (Table 11) and the film thickness reported by Schwarzenbach et al. of 200 pm (12) and Mackay and Yuen (27) produces still-air values of k, of 0.18 and 0.012 m/day, respectively. The lowest wind speed observed in Green Bay of 1 m/s gave a k, of 0.014 m/day using eq 3. The gas film-transfer coefficient has received much less attention since the resistance to transfer for most compounds currently of interest occurs almost entirely in the liquid film. Water is a good example of a substance whose 78
diffusivity in water in air (cmz/s, ~ 1 0 “ ) (cm2/s)
Envlron. Sci. Technol., Vol. 27, No. 1, 1993
4.95 4.70 4.48 4.27 4.10 3.94 3.80
0.059 0.057 0.054 0.052 0.052 0.048 0.047
Schmidt no. in water in air 2300 2425 2550 2665 2780 2890 3000
2.58 2.71 2.83 2.94 3.05 3.15 3.25
exchange is controlled by transfer across the gas film. Many studies have involved studying the rate of evaporation of water itself from a water body and correlating it to the rate of volatilization of the compound of interest. Schwarzenbach et al. (12)took correlations obtained by several different investigators and deduced a suitable approximation for ka(Hzo): ka(HzO) = 0.2~10+ 0.3 (7) where ka(Hzo)is in units of centimeters per second. They as follows: related ka(H1o)to ka(org ka(org chem)
-- ka(H,O) LDa(org
~hem)/~a(HzO)]~
(8)
where Da is the molecular diffusivity in air and n lies between 0.5 and 1 (35). Therefore, the transfer of water vapor across the gas film is calculated as a function of wind speed and then its transfer coefficient is used to calculate the transfer coefficient for some other chemical. Smith et al. (36)estimated n as 0.61. The molecular diffusivities of water vapor and PCBs through air were calculated using the method of Fuller et al. (34). The diffusion coefficients were then used in the above equation to calculate the transfer coefficient of PCBs in air. The average calculated diffusivities and Schmidt numbers in air and water for various PCB homolog groups are listed in Table 11. Henry’s law constant describes the equilibrium partitioning between a liquid and gas phase. Accurate H values are crucial for determining the magnitude and direction of transport of chemicals between air and water in the environment. The H values used in the flux calculations here were obtained from Brunner et al. (37),who reported values for 58 congeners from a technical PCB mixture by using a dynamic method involving a column operating in the concurrent mode. They used their measured H to establish semiempirical structure-property relationships for the H of PCBs they did not measure. Brunner et al. (37) reported a range of H values for PCBs from 0.4 X loe3 to 12 X atm m3/mol (25 “C)in agreement with previously measured ranges. Henry’s law constants were found to be correlated to molecular weight and degree of ortho chlorine substitution. Their predictive equation for H’ incorporating number of chlorine atoms and number of chlorine atoms in the ortho position was used to calculate H values for congeners they did not measure. A list of the H values used can be found in Table 111. Henry’s law constants [from Brunner et al. (37) at 25 “C] were corrected for the appropriate water temperature using a factor calculated by the correlation of Tateya et al. (39) for the relationship of H and temperature. The ratio of Henry’s law constants calculated at 25 “C (298 K) and at the actual water temperature as used to correct the H of Brunner et al. at 25 “C to the actual field temperature: log H,,,/log HT = (7.91 - 3.414/298)/(7.91 - 3414/T) (9)
Table 111. Henry's Law Constantsn
log Hn IUPAC congener no. (atm.m3/ mol) 4, 10 5, 8 6 7 12, 13 16, 32 17 18 19 22 24, 27 25 26 28, 31 29 31, 28 33, 53, 21 37,42 40 41, 71, 64 44 45 46 47,48 49 51 52,43 56,60 63 66,95 70, 76 74 82 83 85 87,81 91 92, 84, 89 97 99 100 101
-3.64 -3.64 -3.60 -3.85 -3.31 -3.70 -3.59 -3.60 -2.64 -3.85 -3.66 -3.77 -3.77 -3.70 -3.70 -3.72 -3.77 -4.00 -4.00 -3.85 -3.91 -3.91 -3.73 -3.72 -3.68 -3.73 -3.70 -4.09 -4.09 -3.92 -4.00 -4.00 -4.23 -4.23 -4.18 -4.23 -4.05 -4.23 -4.13 -4.11 -4.05 -4.05
IUPAC congener no.
log Ha (atmm3/mol)
107 110,77 118 119 128 131 134, 114 135, 144, 147,124 136 141 146 149 151 153, 132, 105 158 163, 138 170, 190 172, 197 173 174 175 176, 137, 130 177 178, 1'29 180 183 185 187, 182 189 191 193 194 195,208 198 199 200, 157 201 202, 171, 156 203, 196 205 206 207
-4.41 -4.23 -4.41 -4.23 -4.89 -4.41 -4.31 -4.25 -4.06 -4.19 -4.60 -4.37 -4.23 -4.62 -4.55 -4.82 -5.05 -4.89 -4.85 -4.85 -4.69 -4.51 -4.69 -4.64 -5.00
-4.69 -4.80 -4.69 -5.05 -4.87 -4.87 -5.00 -4.96 -4.85 -5.00 -4.83 -5.01 -4.83 -5.01 -5.19 -5.33 -5.15
nSource, Brunner et al. (37). Henry's law constants include those experimentally determined by Brunner et al. and others that were calculated using their structure-property relationships.
where Tis the absolute temperature of the water. Henry's law constants decreased by a factor of 2.5-3.2 for each 10 "C decrease in water temperature. This has the net effect of decreasing volatilization fluxes and perhaps reversing the direction of PCB transfer at low temperatures. The magnitude of flux of a chemical across the air-water interface is highly dependent on wind speed; therefore, a description of the wind speed measurements used in the flux calculations follows. Wind speed measurements were taken hourly from the bow of the RV Bluewater at 3 m above the water surface during the period of air sampling. The wind speeds were corrected to 10 m for use in the flux equations by means of the following correlation (12): u, = ((ln z + 8.1)/10.4)u10 (10) where y is the wind speed measured at distance z (m) from the water surface and a boundary roughness height of 0.03 cm is assumed. Wind speed measurements taken at three land sites (Green Bay master site, Fayette State Park, Peninsula State Park) on the eastern side of the bay were obtained (courtesy of C. Sweet, Illinois State Water Survey) for comparison to our overwater wind speed measurements. The land measurements were comparable to our overview measurements. The wind speeds used in the
flux calculations are the median values of the wind speeds we measured during the sampling period at 3 m over the water and corrected to 10 m. Results and Discussion Analysis of bulk water and air for PCBs yields a particulate and a nonparticulate fraction, the sum of which is the total PCB concentration. The concentration of gas-phase PCBs for the 14 days for which volatilization rates were calculated ranged from 0.25 to 2.3 ng/m3. Particle-phase PCBs were below the detection limit in all cases. Greater than 90% of the total PCBs in the atmosphere is in the gas phase (e.g., refs 11,13,40, and 41). We found that location in the bay, i.e., distance north from the mouth of the Fox River, determines the concentrations of PCBs over Green Bay. Concentrations over the bay decrease with increasing distance north in the bay. Several investigators have suggested that PCB air concentration is directly related to air temperature (13, 40-43). The range of air temperatures during the three sampling periods in June, July-August, and October was 2-32 "C. A plot of air concentration vs air temperature failed to exhibit a significant relationship;air concentration differences over the bay could only be related to sampling location. The range of values seen at all sites during the three sampling trips compares well with other measurements taken in the Great Lakes area. Manchester-Neesvig and Andren (41) measured average summer concentrations of 1.8 ng/m3 at a remote site in northern Wisconsin. Swackhamer et al. (13) found slightly higher average summer concentrations of 2.8 ng/m3 for Isle Royale, Lake Superior. Eisenreich (11)reported that air collected over Lake Superior during the summer from 1978 to 1981 and found an average PCB concentration of 1 ng/m3. Baker and Eisenreich (5) measured PCB air concentrations over Lake Superior in 1986 and found an average of 1.20 ng/m3. They concluded that atmospheric PCB concentrations have remained relatively constant over the northern Great Lakes during the last 10 years despite lower PCB loadings. An average of 75 PCB congeners were identified in gas-phase samples. The less chlorinated congeners were enriched in the gas phase in agreement with others (5,11,13,41). The gasphase congener distribution was very similar to the water (dissolved-phase) congener distribution. Figures 3 and 4 show congener patterns in air and water for samples from site 4 and site 21 in July-August, and Table IV lists the congener order for the figures. The congener patterns in air and water at site 4 (high concentrations) are more similar than patterns at site 21 (low concentrations). Both sites have fewer congeners in the dissolved phase in the water compared to the vapor phase in the air; however, all samples are dominated by di- to tetrachloro-substituted congeners. The concentration of total PCBs in Green Bay surface water samples ranged from 0.59 to 41.5 ng/L. Dissolved PCB concentrations ranged from 0.46 to 8.0 ng/L. Ranges for air concentrations (vapor phase) and water (dissolved phase) are shown in Figure 5. PCB concentrations are highest in the lower bay and decrease with increasing distance northward from the mouth of the Fox River for air and water samples, in agreement with the results of Swackhamer and Armstrong (2). Total PCB concentrations at sites 4 and 10, in the southern section of the bay, were 15.6 f 11.2 ng/L ( N = 16), considerably higher than the rest of the bay. Total PCB concentrations at site 14 were intermediate, 5.5 f 2.4 ng/L (N = 5). Measurements by Swackhamer and Eisenreich (unpublished data) of 4.0 ng/L at approximately the same location in 1987 agree well with our values. Total PCB concentrations for the upper Envlron. Sci. Technol., Vol. 27, No. 1, 1993 70
Air 10.0 I P C B = 0.25 ng/rn3
n c
I P C B = 2 . 3 ng/rn3
h
n
0a-r
-
-
I
I
!
I
I
I
I
I
I
I
I
I
I
I
I
Wcter
Water
2.0 ZPCB = 0.35 ng/L
0
I
ZPCB = 4.0 n g / L
01
I
1.5
0
0
c
2.0
v
E
1.0
M
E
1 '.
E"
1 '.
E
0.5 0.0
-d -
-
o b m t - a - o m w ~ o h r n v - r . N t tn h - m - o M r) o, M m m b o O r . - - - - - - - - - N b a 3 0-
6 6 rnbb 7
-
v
-
6
0 N
1 .o
0.0
Congener Number
Figure 3. Congener distribution profiles for air (vapor phase) and water (dissolved phase) for samples from site 21, July 29, 1989.
Table IV. IUPAC Congener Numbers for Figures 3 and 4 4, 10 6 12, 13 16, 32 17 18 26 27, 24 28, 31 29 31, 28 33, 53, 21 37,42 40 44 45 46 47,48 49 51 52,43 56,60 63 66,95 70, 76 74 82
83 85 87, 81 91 92, 84, 89 97 99 100 101 107 110,77 118 119 128 131 134, 114 135, 144, 147, 124 136 141 146 149 151 153, 132, 105 158 163, 138 167
169 170, 190 172, 197 173 174 175 176, 137, 130 177 178, 129 180 183 185 187, 182 189 191 193 194 195, 208 199 200, 157 201 202, 171, 156 203, 196 205 206 207
bay, sites 18 and 21, were 1.7 f 0.6 ng/L ( N = lo), reminiscient of open Lake Michigan water. Swackhamer and Armstrong (2) reported open Lake Michigan water concentrations of 1.2 f 0.46 ng/L. Total PCB concentrations vary with season, distance northward from the city of Green Bay and the mouth of the Fox River, and TSM concentrations. The large range of total PCB concentra80
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Environ. Sci. Technoi., Vol. 27, No. 1, 1993
Congener Number
Flgure 4. Congener distribution profiles for air (vapor phase) and water (dissolved phase) for samples from site 4, August 1, 1989. I
4
S
10
14
Slte
21
"
+N
Figure 5. Concentration ranges for PCBs In water and air of Green Bay. The concentrations are greatest in the southern bay.
tions (0.6-41.5 ng/L) is due mainly to the variation in the particulate fractions, which was directly dependent on the TSM concentration. The concentration range of PCBs in the particulate fraction in all samples was 730-1300 ng/g. The rather constant concentration on particles implies that PCBs are partitioning to organic carbon and differences in partitioning are due to dilution by inorganic particles. The organic carbon content of the particles ranged from 11 to 37% with an average of 23%. TSM concentrations in surface waters at sites 4 and 10 range from 2.9 to 26.0 ng/L. By comparison, TSM concentrations ranged from 0.76 to 3.3 mg/L for sites and 18 and 21, similar to open Lake Michigan TSM concentrations of 0.5-2 mg/L (44). Generally 60-75 congeners were identified in the dissolved (XAD-2) phase and 75-85 congeners in the particulate (filter) phase of water. The congener profiles of both the particulate and dissolved phases are dominated by tri- and
Site 4, 6 - 1 0 - 8 9
Site 18, 6 - 5 - 8 9 7.5 LogKoc = .23LogKow f 4 . 6 2 r7-=.43
.
7.0 6.5 6.0
- - a 5.5 0
0
D
+@ .
5.0
Y 0 0 1
Site 18, 10-20-89
Site 4, 1 0 - 2 3 - 8 9
7.5 LogKoc=.l8LogKow+4.79 r2=.46 7.0
.
6.5
.
LogKoc=.26LogKow+4.49 r2=.66
1
6.0 0.
5.5 1
5.0 5.0
5.5
. . 6.0
6.5
1 7.0
7.5
8.0
5.0
5.5
6.0
6.5
7.0
7.5
8.0
Log K o w Figure 6. log KO,vs log K, for two samples from site 18 and two from site 4. The slope of the lines is much less than the hypc ... stlcal value of 1 in all cases.
tetrachlorinated congeners and are similar in both phases. Swackhamer and Skoglund (45) suggested that the lower chlorinated congeners in the dissolved phase partition into phytoplankton, which dominated the particulate phase in Green Bay during our sampling periods. Volatilization is mediated by the partitioning of PCBs between water and aquatic particles suspended in solution because only truly dissolved PCBs are available for transfer across the air-water interface. Partitioning is described by a distribution coefficient, Kd (L/kg): Kd = (C,/TSM)/(Cd) (11) where C, is the concentration in the operationally defiied particulate phase (ng/L), TSM is the concentration of TSM (kg/L), and Cd is the concentration of PCBs in the operationally defined dissolved phase (ng/L). The distribution coefficient describes the observed distribution between the operationally defined particulate and dissolved phases whereas Kp,the partition coefficient, is the true thermodynamic equilibrium coefficient (46). The range of observed distribution coefficients for total PCB samples throughout the bay was relatively small, 1.1 X 106-4.6 X lo5L/kg, much smaller than the range observed by Swackhamer and Armstrong (2) of 8.0 X 10'3.3 X lo6 L/kg. Baker et al. (46) summarized PCB congener distribution coefficients measured in the Great Lakes and elsewhere; they reported an overall average log Kd = 5.5 for total PCBs compared to an overall average log Kd = 5.4 for all congeners in all samples taken in Green Bay. The partitioning of PCBs to aquatic particles is expected to increase as the hydrophobicity of the compound in-
creases (47-50). However, the distribution of PCBs between dissolved and particulate phases is not related strongly to KO,in this case. The organic carbon content (OC) of the solid phase is an important factor affecting partitioning in laboratory experiments, so the distribution coefficient can be expressed in terms of the OC content of the sorbent as follows:
K , = &/oc (12) where OC represents the fraction of the total solid that is organic carbon. Figure 6 shows the relationship between log KO,and log KO,for individual congeners in samples from sites 4 and 18 in June and October. The slope of the K , vs Kowrelationship should be 1 according to theory and practice (46, 48,501, but our data from Green Bay show that the slope log K , vs log KO,is -0.2 and less than the hypothetical value of 1. Other investigators have observed this discrepancy between field data and laboratory correlations (22, 46, 51) and have attributed it to the presence of a third phase or colloids as described by Gschwend and Wu (52). Organic-rich colloids pass through the glass fiber filter and may contribute to the dissolvedphase concentration. Therefore the dissolved phase may be overestimated and cause Kd to decrease with increasing solids concentration. The PCBs associated with colloids in surface waters are not directly available for exchange across the air-water interface; therefore, the importance of colloids in a given system needs to be examined. The phenomenon of decreased distribution coefficients with increasing solids concentrations (e.g., refs 15, 49,53, and 54)is thought to be a result of increased colloidal-phase
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Envlron. Sci. Technol., Vol. 27, No. 1, 1993 81
Table V. PCB Concentrations in Air and Water and Water of Green Bay Used in Flux Calculations
site
date
18 18 10
6-4-89 6-5-89 6-6-89 6-7-89 6-10-89 6-11-89 7-28-89 7-29-89 7-30-89 7-31-89 8-1-89 10-21-89 10-22-89 10-23-89
io 4
io
18 21 14 10 4 14 10 4
[PCBI (ng/m3) air water (vapor) (dissolved) 0.53 0.36 1.04 1.58 1.07 1.23 0.31 0.25 0.41 0.70 2.29 0.75 1.25 1.38
1080 767 1830 1500 7800 1830 1170 350 1520 5860 4050 2450 6670 2240
material with increasing suspended solids. Our data show only a 3-10-fold decrease in Kd as the TSM concentration increased from 0.8 to 26 mg/L. Baker and Eisenreich (5) found that correcting for colloid binding affected the calculated fluxes of the most soluble PCBs (which contribute the most to a volatilization flux) by only a small amount. Eadie et al. (55) reported that binding of HOC to DOC in Green Bay water is small, and on average, the amount associated with this phase is less than 10%. Correcting the data for colloids in this study is not defensible. Another possible explanation for the apparent lack of increased sorption with compound hydrophobicity is that sorption-desorption kinetics are slow compared to particle generation rates. Phytoplankton dominated the particle population in Green Bay (45). The mechanism for phytoplankton uptake of PCBs is analogous to partitioning of PCBs to sediment material. A two-step mechanism involving initial, rapid surface sorption followed by a much slower diffusion into the aggregate, in this case the lipid complex of the cell, has been proposed by Swackhamer and Skoglund (45) to describe the partitioning of HOCs to algae. Swackhamer and Skoglund have shown that the rate of uptake of PCB congeners by phytoplankton is slow relative to algal growth rates during summer, and therefore, PCBs partitioning to algae do not achieve equilibrium. Their laboratory studies and field data from Green Bay show that it is unlikely that PCBs in the dissolved phase reach equilibrium with phytoplankton during normal
growth periods. Their results from winter (low growth rates) showed a very strong relationship of partitioning to KO,, showing that the PCBs were at equilibrium during low-growth periods (cold temperatures). The lack of increased partitioning with increased hydrophobicity can be explained by their results, which indicate that PCBs had not reached equilibrium with phytoplankton from Green Bay during the summer of 1989. We have concluded that the dissolved-phase concentration of PCBs as it was measured best reflects the quantity available for air-water exchange.
Volatilization Fluxes Air-water transfer rates were calculated for 14 days spanning three cruises in Green Bay in 1989 by using the stagnant two-film model: flux = Kol(C, - C*) (13) where the net flux equals the overall mass-transfer coefficient multiplied by the concentration gradient. Calculated Ko
15.0
I
-
v
5
10.0
, , , , , , , I
, , ,
- 5 . 0 ) , 8.0
r c o n g e n e r s = 30 ng/m2/doy
-2.0
j
= 1
I 1
1
1
1
1
1
1
1
1
1
1
1
1
- - -
,
N
Congener Number
Flgwe 8. PCB congener flux patterns at site 4 on June 10, 1989 (top) and October 23, 1989 (bottom).
10 m (ul0). The overall mass-transfer coefficient is controlled by ul0 at very low wind speeds. This is demonstrated by the events labeled A and B in Figure 9. The calculated flux of total PCBs on July 31 at site 4 (A) with a low wind speed of 1 m/s is 90 ng/m2.day. On August 1, the following day (B), water concentrations and air and water temperatures were comparable to those on July 31 (refer to Table VI) but the wind speed increased to 4 m/s.
The result of the higher wind speed was remarkable; the calculated flux increased to 800 ng/m2.day, nearly 9 times larger than the previous day. The combined influence of C, and ul0 may be demonstrated by selecting events when both were high values, and an event where one was low and one was high. For the event labeled C in July (site 21), the ul0 at 5.5 m/s was at the high end of the range while C, at 350 ng/m3 was the lowest concentration observed. This combination yielded a total PCB flux of 70 ng/m2.day. In comparison, C, (-6670 ng/m3) and ul0(-6.5 m/s) at the event labeled D in October (site 10) were both high and yielded the highest calculated total PCB flux of 1300 ng/m2-day. The water concentration in Green Bay drives the volatilization out of the water and wind speed determines the masstransfer coefficient. Another factor important in the determination of PCB flux across the air-water interface is water temperature. Temperature exerts its influence by affecting the magnitude of H. Previously we noted (eq 9) that H decreases by a factor of 2.5-3.2 for every 10 "C change in temperature. Henry's law constant affects both the water concentration of PCBs in equilibrium with the measured air concentration (C* = P / H , eq 1) and the magnitude of the gas film mass-transfer coefficient (Hk,/RT, eq 2). The relative influence of wind speed and water temperature on total PCB flux is demonstrated in Figures 10 and 11 for samples from site 21 and site 4 in July-August, 1989 for which congener distribution patterns are shown in Figures 3 and 4. The calculations were made for wind speeds of 2-8 m/s and for water temperatures of 5-15 "C. These values are typical of conditions observed in Green Bay. For each wind speed a total PCB flux was calculated using H appropriate for the water temperatures. The PCB flux varied from
