Air−Surface Exchange of Polybrominated Diphenyl Ethers and

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Environ. Sci. Technol. 2002, 36, 1426-1434

Air-Surface Exchange of Polybrominated Diphenyl Ethers and Polychlorinated Biphenyls T. GOUIN,† G. O. THOMAS,‡ I. COUSINS,† J. BARBER,‡ D . M A C K A Y , * ,† A N D K . C . J O N E S ‡ Canadian Environmental Modelling Centre, Trent University, Peterborough, Ontario, K9J 7B8, Canada, and Environmental Science Department, Institute of Environmental and Natural Sciences (IENS), Lancaster University, Lancaster, LA1 4YQ, UK

Air and leaf-litter samples were collected from a rural site in southern Ontario under meteorologically stable conditions in the early spring, prior to bud burst, over a threeday period to measure the simultaneous diurnal variations in polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs). PBDEs are used in a wide range of commercial products as flame retardants and are being assessed internationally as potential persistent organic pollutants. Total PBDE concentrations in the air ranged between 88 and 1250 pg m-3, and were dominated primarily by the lighter congeners PBDEs 17, 28, and 47, and concentrations of total PCBs ranged between 96 and 950 pg m-3, and were dominated by the lower chlorinated (tri- to tetra-) congeners. Slopes of Clausius-Clapeyron plots indicate that both PCBs and PBDEs are experiencing active air-surface exchange. Fugacities were estimated from concentrations in the air and leaf-litter and suggest near equilibrium conditions. Following the three-day intensive sampling period, 40 air samples were collected at 24hour intervals in an attempt to evaluate the effect of bud burst on atmospheric concentrations. Total PBDE concentrations in the daily air samples ranged between 10 and 230 pg m-3, and were dominated by the lighter congeners PBDE 17, 28, and 47, whereas concentrations of total PCBs ranged between 30 and 450 pg m-3 during this period. It is hypothesized that the high PBDE concentrations observed at the beginning of the sampling period are the result of an “early spring pulse” in which PBDEs deposited in the snowpack over the winter are released with snowmelt, resulting in elevated concentrations in the surface and air. Later in the sampling period, following bud burst, PBDE concentrations in air fell to 10 to 20 pg m-3, possibly due to the high sorption capacity of this freshly emerging foliage compartment.

Introduction There are several reports of the seasonal temperature dependence of atmospheric concentrations of semi-volatile organic compounds (SVOCs), such as polychlorinated biphenyls (PCBs) and organochlorine pesticides, on a relatively * Corresponding author phone: (705) 748-1489; fax: (705) 7481569; e-mail: [email protected]. † Canadian Environmental Modelling Centre. ‡ IENS. 1426

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long-term basis (1-6). Because of their moderate vapor pressure, low solubility, and low reactivity, these compounds exchange between the atmosphere and various environmental surfaces, such as soil, vegetation, and water, as a function of temperature in a series of volatilization/sorption “hops”, commonly referred to as the “grasshopper effect” (7). It is believed that the temperature dependence of these air-surface exchange processes is a key factor in understanding the distribution and movement of SVOCs far from areas where they are discharged (1, 2, 8, 9). This phenomenon, combined with their persistence, implies that SVOCs, including substances no longer produced or used, can reach and are still detectable in remote areas. Fewer studies have examined how SVOCs cycle in response to diurnal temperature changes. It has been reported that atmospheric concentrations of PCBs, polyaromatic hydrocarbons, polychlorinated naphthalenes, and some organochlorine pesticides exhibit diurnal variability (8, 10-12), and Hung et al. (13) have recently reported diurnal variation of PCBs in grass samples. Because it is not yet possible to directly measure actual fluxes of SVOCs at ambient concentrations with such a short time resolution, data on concentration cycling can be interpreted to deduce rates of transfer and directions, thus leading to an improved understanding of the environmental cycling of these substances. Hoff et al. (3) first suggested interpreting such data using Clausius-Clapeyron (CC) plots, in which the natural logarithm of the partial pressure of the gas (P) is regressed with reciprocal absolute temperature (1/T), yielding slopes which correspond to enthalpies similar to reported heats of vaporization. They concluded that volatilization from surfaces relatively near the sampling location was controlling the air concentrations (3). The slopes were observed to flatten out at temperatures below freezing, possibly because of the increasing importance of long-range atmospheric transport (LRAT) relative to local volatilization, the relative strength of the sources in the various environmental media volatilizing material to the air, and changes in the nature of the exchange process with changing seasons and temperatures (1). Thus, by comparing the slopes derived from CC plots to known enthalpies of air-surface transfer, insights can be gained about the relative importance of local and distant sources (14). Vegetation is a potentially important environmental compartment mediating the diurnal cycling of SVOCs in the atmosphere. Thomas et al. (15) suggested that foliage growth exposes fresh surfaces for sorption of SVOCs, which could be sufficient to lower air concentrations when plant growth is most profuse. This phenomenon would be most evident in the spring, when the growth of foliage is rapid. Wania and McLachlan (16) have recently investigated this “springbuffering” effect by using a multimedia model to demonstrate the importance of a rapidly growing vegetation compartment. In this study, the simultaneous diurnal variation of concentrations of polybrominated diphenyl ethers (PBDEs) and PCBs prior to bud-burst is examined. PBDEs are in a class of flame-retardants that have been identified as persistent organic pollutants. They are used in a wide range of consumer products, including plastics, electronic devices, and textiles (17). Total global consumption of these compounds has greatly increased over the past twenty years with an estimated global production of 50 000 tonnes per year, being distributed primarily between North America (40%), the Far East (30%), and Europe (25%), during the 1990s (17, 18). Because of their potential for bioaccumulation in aquatic food chains and presence in environmental samples of air, 10.1021/es011105k CCC: $22.00

 2002 American Chemical Society Published on Web 03/03/2002

water, sediment, and fish in areas far from where they are produced, they are of growing environmental concern (1720). Compared to PCBs, relatively little information exists on environmental levels of PBDEs, especially pertaining to atmospheric and temperature-dependent data. Thus, by comparing the diurnal data obtained for PCBs, which have been well studied, to data obtained for PBDEs for the same samples, we hope to provide insights about the air-surface exchange of these compounds and determine if these substances behave similarly.

Materials and Methods The sampling site, shown in Figure S1, is located near Pigeon Lake, at the James McLean Oliver Ecological Centre (JMOEC) (44° 33′ N, 78° 30′ W), which is about 25 km NNW of Peterborough, Ontario, Canada. The 110-hectare rural waterfront property consists of lake shoreline, pastures, marshland, and forested habitat. The research site is representative of the mostly rural nature of southern Ontario and has no industrial sources within 25 km. Meteorological conditions over the period were monitored and recorded both at the field site and at a fully equipped meteorological station located at Trent University in Peterborough. Thirty-six air samples were taken at 2-hour intervals, and 12 leaf-litter samples were collected at 6-hour intervals over an intensive three day sampling period in early spring, prior to bud burst, to evaluate diurnal variations without an active vegetation compartment. Air samples were then taken at 24-hour intervals over the next 40 days to evaluate the effect of bud burst on atmospheric concentrations. Two highvolume air samplers (General Metal Works model GPS1 PUF sampler) were placed, 1.5 m above the soil surface, within a young hardwood forest consisting mostly of Sugar Maple (Acer saccharum) and Black Ash (Fraxinus nigra). Sampling began at noon on April 24, 2000, with samples being collected serially at 2-hour intervals over a three-day period, finishing at noon on April 27, 2000. The two high-volume air samplers were run simultaneously, with the samples being combined, resulting in bulked sample volumes of 80-100 m3. The particulate fraction was trapped on a Whatman glass microfiber filter (GMF) (grade GF/A, 10-cm diameter to retain particles above 1.6 µm in diameter) and the gas-phase fraction was trapped on a polyurethane foam plug (PUF) (length 8 cm, diameter 6.25 cm, density 0.035 g cm-3). The GMFs were pre-cleaned by baking in a furnace at 450 °C for 24 h and then stored in solvent-rinsed aluminum packages until use. The PUF plugs were pre-extracted in dichloromethane (DCM) for 18 h on a Soxhlet system; the DCM was removed by placing the PUF plugs in a solvent-cleaned desiccator under vacuum. The plugs were then stored in solvent-rinsed glass jars with aluminum-lined lids. Clean PUF plugs and GMFs were loaded, using solvent-rinsed tongs, into sampler modules prior to sampling. After sampling, the sample modules were unloaded and immediately rinsed with solvent, and the PUF plugs and GMFs were replaced in the original sample jars and frozen at < -5 °C. Field blanks were also prepared and handled in a manner identical to that of the samples, i.e., the plugs and filters were loaded into and unloaded from the sampler in a manner identical to that of the samples, but no air was aspirated through them. Leaf-litter samples, consisting of approximately 50 g of randomly sampled partly decomposed sugar maple leaves, were collected from the litter on the forest floor in the vicinity of the air samplers at 6-hour intervals. They were placed into bags in the field, immediately sealed, and stored at < -5 °C until they were extracted in the laboratory at Lancaster University, U.K. The GMFs and PUF plugs were combined and spiked with labeled recovery standards (13C PCBs 28, 52, 101, 153, 138, 180, and 209) and Soxhlet extracted for 18 h with 250 mL of DCM. The extracts were reduced under nitrogen to

500 µL and placed on a silica fractionation column with a minimum of washings (3 × 0.5 mL). The compounds of interest were eluted with hexane (32.5 mL) then 1:1 hexane/ DCM (15 mL), collected together as one fraction, then reduced under nitrogen to 500 µL and placed on the Biobeads SX-3 GPC column using a minimum of washings (3 × 0.5 mL). Two 15-mL fractions were eluted from the GPC column with 1:1 hexane/DCM. The first fraction was discarded and the second fraction, containing the PBDEs and PCBs, was solventexchanged into dodecane. PCB 30, 13C PCB 141, and 13C PCB 208 were added as internal standards, and the sample was reduced under nitrogen to 25 µL for injection on the GCMS. Five of the leaf-litter samples, corresponding to collection at 12-hour intervals (at the approximate maximum and minimum temperatures), were selected for analysis, because it is believed that the concentration of PCBs and PBDEs at the terrestrial surface would not be significantly affected by the diurnal atmospheric concentration changes during the study. Leaf-litter samples were prepared and analyzed by grinding a 25-g sample with anhydrous sodium sulfate and spiking it with 13C-labeled PCB recovery standards. Samples were then Soxhlet-extracted with 250 mL of DCM for 18 h. The volume of the extract was reduced under nitrogen to 500 µL, and the sample was cleaned up using a method reported by Thomas et al. (21). The internal standards used for the PBDE analysis were 13C PCB 141 and 13C PCB 208. Method blanks were prepared following the same procedure. Organic matter (OM) contents of leaf-litter samples were determined by ashing the dried material at 450 °C overnight. All PBDE analyses were undertaken with a Thermo Trace GC-MS with separation on a 0.18-mm i.d. DB5 MS 30-m column. The mass spectrometer was run with an NCl source in SIM mode, using ammonia as the reagent gas. A total of 21 PBDEs were quantified using an internal standards method. All PCB analyses were undertaken by GC-MS on a Fisons MD800 running with an EI+ source in SIM mode. Separation was achieved on a 0.25-mm i.d. CP-Sil 8 50-m column. Samples were run in splitless mode with helium as the carrier gas. A total of 47 PCB congeners were quantified using an internal standards method. The method detection limit (MDL), defined as 3 × standard deviation of the blank values, expressed in pg m-3, for the PCBs ranged between 0.2 and 37.5 pg m-3, and for the PBDEs ranged between 0.3 and 47 pg m-3. Full details of the MDLs and blank values for the PCBs and PBDEs are given in Table S1 in the Supporting Information. The recoveries for the labeled PCB recovery standards were consistently >92%. Concentrations of PCBs and PBDEs were blank corrected, but not corrected for the recoveries of labeled compounds.

Results Air Concentrations: Intensive Sampling Period. Meteorological conditions over the three-day intensive sampling period (summarized in Table S2, Supporting Information) show a very large diurnal fluctuation in temperature, ranging between -2 and 18 °C, which is typical for the time of year. There was no precipitation, and winds were predominantly from the north, under relatively stable conditions, with speeds from 0.1 to 2.5 m/s. The ΣPCB (sum of the 47 congeners) concentrations in the air ranged between 96 and 950 pg m-3, (Table S1) which is consistent with rural concentrations measured in southern Ontario (1). The ΣPCB concentrations was dominated by the lower chlorinated (tri- to tetra-) congeners, PCBs 18, 22, 28, 31, 44, 49, and 52 (concentration means and ranges are given in Table S1). Of the heavier congeners, only PCBs 87, 99, 110, and 118 were consistently detected at concentrations above the MDL. The ΣPBDE (sum of the 21 congeners) concentrations in the air ranged between VOL. 36, NO. 7, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 1. Atmospheric concentrations and ambient temperatures observed during the 3-day intensive sampling period for (a) PCB congeners 18, 28, 52, and 118, and (b) PBDE 47 and ΣPBDE. 88 and 1300 pg m-3, and were dominated primarily by the lighter congeners, PBDEs 17, 28, and 47. The concentrations of individual PCB and PBDE congeners followed similar trends, as shown in Figure 1. The 5 data points missing from Figure 1a and the 4 data points missing from Figure 1b were from samples lost during cleanup. 1428

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Air Concentrations: Daily Samples. The ΣPCB concentrations measured in air samples taken on a daily basis over the duration of a month, following the 3-day intensive sampling period, ranged between 30 and 450 pg m-3. The concentration profiles of selected PCB congeners in air are shown in Figure 2a, where air concentrations for April 2427, 2000 are representative of a 24-hour time-averaged

FIGURE 2. Daily concentrations observed between April 24, 2000 and May 25, 2000 for (a) PCB congeners 18, 28, 52, and 118, and (b) PBDE 47 and ΣPBDE. concentration, derived from the concentration data collected during the intensive sampling period. The ΣPBDE concentrations measured in the daily air samples ranged between 10 and 230 pg m-3, and were dominated by PBDE congeners 17, 28, 47, and 99. Figure 2b shows the concentration profile observed for ΣPBDE and PBDE 47 over the daily sampling

period (April 28 - May 25), whereas concentrations for April 24-27, 2000 are representative of the 24-hour time-averaged concentration. The concentrations reported above for ΣPCB and ΣPBDE during the daily sampling period, and represented in Figures 2a and 2b respectively, are for only those samples that were analyzed during that period. Because the collection VOL. 36, NO. 7, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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of samples on a daily basis was intended to evaluate the effect of bud burst on atmospheric concentrations, five 24hour samples from between May 02 and May 08, 2000 were selected for analysis. This is the period directly following the first observance of bud burst of sugar maple. A number of samples were then randomly selected over the remaining period to establish any significant trend. Leaf-Litter Concentrations: Intensive Sampling Period. The organic matter content of the leaf-litter samples averaged 81 ( 0.8% of dry matter. The ΣPCB concentrations in the leaf-litter samples ranged between 5 and 7 ng/g DM, and are typical of concentrations found in grass samples in rural areas (13). The ΣPBDE concentrations in leaf-litter samples ranged between 5 and 14 ng/g DM, and were dominated by PBDEs 47 and 99. All of the 21 PBDE congeners were detected except for PBDEs 35, 77, 166, 181, and 190 (Table S3). Although the concentrations of ΣPCBs at the terrestrial surface are in accord with previously reported levels, there appear to have been no other studies of PBDEs at the terrestrial surface.

Discussion The ΣPBDE concentrations in the air and leaf-litter samples collected during the 3-day intensive sampling period are greater than the ΣPCB concentrations, which differs from observations in other studies. For instance, Strandberg et al. have reported ΣPBDE concentrations in rural ambient air ranging between 5 and 20 pg m-3, and observed concentrations as high as 80 pg m-3 in urban air taken from Chicago (22). They reported ΣPCB concentrations in rural ambient air that were much higher than their ΣPBDE concentrations, ranging between 100 and 1200 pg m-3, and observed concentrations as high as 4100 pg m-3 in the samples from urban Chicago (22). Although the relatively high levels of PBDEs in the air observed during the intensive sampling period of this study greatly exceed the levels reported by Strandberg et al., they are comparable with data for archived air samples collected from Tagish, Yukon between 1994 and 1995, that ranged between 10 and 750 pg m-3 (23). Figure S1 shows two-day back trajectories for the intensive study period, determined using the HYSPLIT trajectory model available on-line from the NOAA Air Resources Laboratory (24). The air collected during the study period was from the largely unpopulated north, where there are no known sources of PCBs or PBDEs. Considering that the meteorological conditions during the intensive sampling period were relatively stable, the elevated concentrations of PBDEs observed during this period are surprising. We believe that the observed concentrations in air are unlikely to be due to direct advective transport of PBDEs and PCBs from a specific source, but result from a local unidentified source or airsurface exchange. The latter is our preferred explanation, because the concentrations of PCBs and PBDEs both followed the diurnal fluctuations observed in ambient air temperatures (Figure 1). The correlation between temperature and atmospheric concentration is significant for PCB 18, which has a calculated Pearson correlation coefficient, r, of 0.39 and a p-value of 0.035. The correlations for PCBs 28, 52, and 118 are not significant, having calculated p-values of 0.47, 0.15, and 0.74, respectively. The diurnal fluctuation of atmospheric concentrations of PBDEs is more pronounced than that observed for the PCBs, as shown by correlation coefficients of r of 0.67, 0.70, and 0.54 for PBDEs 17, 28, and 47 respectively, and by p-values < 0.01 for each of these congeners. It may be that the PBDEs displayed a stronger correlation with temperature during this period due to the manner in which they are sorbed to the terrestrial surface. Hornbuckle and Eisenreich (11) suggest that the strong diurnal variations of SVOCs observed in air are most likely dominated by surface adsorption processes, because adsorption response times to changes in 1430

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temperature are relatively fast, on the order of seconds, whereas absorption response times are on the order of days. Goss and Schwarzenbach (25) have similarly demonstrated that adsorption has an important influence on air-surface exchange of SVOCs. One possibility for the difference between PCBs and PBDEs is that because of their greater molar volumes, PBDEs adsorbed at the terrestrial surface diffuse more slowly into the organic matrix and become absorbed more slowly than do PCBs. They thus remain labile and are more susceptible to evaporation. When interpreting the data shown in Figure 1 we have focused on examining the relationship between the air concentration and temperature. A number of other meteorological factors, such as wind speed and direction, and relative humidity, may also be contributing to the observed variability. For instance, Hippelein and McLachlan (26) studied the relative influence of temperature and relative humidity on surface-air partitioning under controlled laboratory conditions, and demonstrated that both factors have a strong effect on the partitioning behavior of SVOCs. However, according to their findings, the large temperature variation observed within the sampling period has a stronger effect on surface-air partitioning than does the variation observed in relative humidity (26). In field studies it is difficult to separate the individual effects of different meteorological parameters, especially because temperature, relative humidity, and wind speed may all co-vary. We hypothesize that the PBDE concentrations observed in air during the intensive sampling period are strongly controlled by temperature, and that these other meteorological factors only play a minor role. Although the ΣPBDE concentrations in air during the intensive sampling period are unusually high, Figure 2b shows that the concentration of PBDEs in air gradually decreased to 10 to 20 pg m-3 over 10 to 14 days, to levels comparable with those reported by Strandberg et al. (22). This observation is consistent with the buffering effect associated with the rapid growth of vegetation suggested by Thomas et al. (15). Bud-burst of sugar maple was observed to occur on May 02, 2000. Thus, as suggested by Thomas et al. (15), the decreasing concentrations in air may be attributable to the sorptive capacity of freshly emerging vegetation. A similar trend, although less pronounced, is observed for the PCBs, shown in Figure 2a, in which concentrations are seen to increase somewhat before decreasing. This increase in atmospheric concentrations coincides with the prevalence of warm southwesterly winds, during the period of May 06-07, 2000 (Table S2), suggesting that the increased concentrations are related to the movement of air from large urban areas in southern Ontario. The effect of this air mass can also be observed for the PBDEs in Figure 2b, where an increase in concentration is also observed for May 06 and 07, 2000. Interpretation of Air Data using Clausius-Clapeyron (CC) Plots. To help interpret these data CC plots were prepared. Figure 3 illustrates the results obtained for PCBs 18, 28, 52, and 118, and for PBDEs 17, 28, and 47. The points to the right and below the arrows (or 1000/(281 K)), which represent air concentrations below 8 °C, show a considerable degree of scatter and do not fit well with the points to the left of the arrow. This observation is consistent with data reported by Hoff et al. (1), who refer to this phenomenon as the “hockey stick” effect. The regression slopes (a), intercepts (b), r2 values, and calculated enthalpies of air-surface exchange, ∆Hcalc, for the points above 281 K for the PCB and PBDE congeners shown in Figures 3a and b, are given in Tables 1 and 2. There is generally good agreement between laboratory measured enthalpies and ∆Hcalc, but it is not possible to assign ∆Hcalc as being either the enthalpy of vaporization, ∆Hvap, or the

FIGURE 3. Clausius-Claperyon plots produced from data collected during the 3-day intensive sampling period, April 24-27, 2000 for (a) PCBs 18, 28, 52, and 118, and (b) PBDEs 17, 28, and 47. The points to the right and below the arrows are air concentration data at temperatures below 8 °C. enthalpy of octanol-air exchange, ∆HOA, as these values are similar. The CC plots for both the PBDEs and PCBs suggest that atmospheric concentrations are being controlled by a combination of factors. At low temperatures, the relatively stable levels of PBDEs and PCBs observed in the air are most

likely being controlled by LRAT, whereas at warmer temperatures the compounds are being driven from the surface to the air due to the temperature dependence of physicalchemical properties, specifically KOA and/or vapor pressure. The indication that PBDEs are experiencing active air-surface exchange during the 3-day intensive sampling period of this VOL. 36, NO. 7, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Summary of Regression Slopes, Intercepts, r 2, and p-values for Data at Temperatures Above and Below 281 K, and Calculated Enthalpies for PCBs 18, 28, 52, and 118 Compared to Enthalpies from Various Other Sources for These Same Congeners PCB congener

a

b

r2

18 22 28 31 44 52 70 74 87 99 110 118

-10 -6.8 -8.3 -7.9 -7.5 -7.9 -5.8 -5.4 -8.1 -8.7 -8.7 -9.6

14.0 0.89 6.85 5.51 3.38 4.81 -2.59 -4.69 3.7 5.6 6.4 8.92

0.57 0.37 0.33 0.50 0.50 0.52 0.25 0.21 0.32 0.42 0.32 0.38

18 22 28 31 44 52 70 74 87 99 110 118

-2.1 -3.5 -2.4 -2.9 -2.7 -2.9 -2.9 -3.1 -5.0 -1.2 -1.2 -2.2

-15.0 -9.94 -13.7 -11.6 -13.1 -12.4 -12.2 -12.2 -6.3 -20.4 -16.7 -16.8

0.04 0.11 0.04 0.11 0.06 0.07 0.05 0.06 0.10 0.007 5.6 × 10-5 0.08

a

From Halsall et al. (2).

b

p-value data >281 K 4.8 × 10-3 3.5 × 10-2 5.1 × 10-2 9.8 × 10-3 9.6 × 10-3 8.6 × 10-3 9.4 × 10-2 0.13 5.4 × 10-2 2.3 × 10-2 5.7 × 10-2 3.2 × 10-2

∆Hcalc (kJ/mol)

∆ Ha (kJ/mol)

∆Hvapb (kJ/mol)

∆HOAc (kJ/mol)

86.5 56.5 69.5 65.7 62.4 65.4 48.2 44.9 67.3 72.3 72.3 79.9

64.7 ( 9.7 n/ad 96.1 ( 13.9 n/a n/a 59.5 ( 8.9 n/a n/a n/a n/a n/a 91.3 ( 7.9

75.3 78.0 78.0 77.7 81.0 80.8 84.8 83.9 87.3 86.8 86.6 89.3

(15) 72.6 n/a (29) 72.6 n/a n/a (53) 45.9 n/a n/a n/a n/a n/a 89.9

data 281 K 9.3 × 10-4 1.6 × 10-3 6.0 × 10-3

17 28 47

-5.90 -7.26 -3.34

3.23 8.92 -2.50

0.59 0.26 0.07

data < 281 K 0.73 0.63 0.10

a

From Tittlemier et al. (31).

b

p-value

CS (0.35focKOWFS/1000H)

(1)

where CS is the concentration (DM) in the leaf-litter, foc is the mass fraction organic carbon, KOW is the octanol-water partition coefficient, FS is the density of the leaf-litter (kg/ m3), and H is the Henry’s Law constant (Pa m3/mol). The coefficient 0.35 was suggested by Seth et al. (33), and is rep1432

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∆Hvapa (kJ/mol)

∆Hvapb (kJ/mol)

∆HOAc (kJ/mol)

117.0 101.4 90.5

n/ad 94.5 103.1

n/a n/a 92.0

72.8 74.5 97

From Wong et al. (29). c From Harner and Shoeib (30).

study suggests that, like the PCBs, PBDEs are also subject to a “grasshopping” effect and consequently have a potential for LRAT. Calculation of Air/Surface Fugacities. To further interpret the air-surface exchange of PBDEs, fugacities of PBDEs were estimated for the air and leaf-litter samples. The fugacity (partial pressure) in air, fA, can be estimated as CART, or CA/ ZA, where CA is the gaseous concentration in air, RT is the gas constant-temperature group, and ZA is the Z-value in air (32). The concepts of fugacity and Z-values are described in detail by Mackay (32) and will not be repeated here. For leaf-litter, the fugacity, fS, was estimated from

fS )

∆Hcalc (kJ/mol)

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d

n/a indicates no data available.

resentative of the relationship between organic carbon and octanol partitioning with respect to water. The leaf-litter contained 81% (DM) organic matter. Organic matter typically contains approximately 50-60% organic carbon (32), thus we have assumed that the organic matter present in the leaflitter is 40% organic carbon. The temperature dependence of KOW and H have been calculated using typical enthalpies for PCBs, as ∆HOW ) -22 kJ mol-1, and ∆HH ) 55 kJ mol-1. Figure 4 shows the estimated fugacities for PBDE 47. It must be appreciated that the estimated fugacities, especially in the leaf-litter, are subject to considerable error, possibly up to a factor of 3 (33). Thus, the absolute fugacities shown in Figure 4 should be regarded as tentative. It is clear, however, that fA and fS follow each other quite closely, and are similar in magnitude, thus near equilibrium conditions apply. This further suggests that the diurnal variation observed in PBDE 47 air concentrations is due to the influence of rapid, localized air-surface exchange processes. Effect of an “Early Spring-Pulse” on Atmospheric Concentrations. We hypothesize that during the winter there is a continuous accumulation of SVOCs in snowfall, by both

FIGURE 4. Fugacities in air over a terrestrial surface for PBDE 47 calculated using concentration data collected in air and leaf-litter samples. aerosol scavenging and sorption to the air-ice surface, as discussed by Wania et al. (34) and Franz and Eisenreich (35, 36). At this site the maximum snow depth reached 0.7 m. During the early spring the snowpack melted, and a fraction of the accumulated PBDEs was transported to the surface litter. Subsequently, this was volatilized back to the overlying atmosphere, generating the pulse of unusually high concentrations observed in the air. Further evidence that supports the suggested importance of snow scavenging are the observations of Blais et al. (37), who report unusually high concentrations of these compounds in snow deposited in the mountains of western Canada. Kelly and Gobas (38) observed unusually high concentrations of PCBs in lichens during the spring, and suggested that this is due to the increased uptake by the lichens of PCBs during the spring melt. Hornbuckle et al. (39) and Hillery et al. (40) have also observed evidence for a spring-pulse of PCBs in air samples collected at Brimley, MI, and Eagle Harbor, MI, where concentrations of ΣPCBs reached an annual maximum of about 900 and 200 pg/m3, respectively, during April and May. It should be noted that, although the present PBDE data are consistent with an early spring-pulse, verification requires more systematic and intensive sampling before, during, and after the alleged pulse. It is notable that the data during the intensive sampling period were obtained prior to bud-burst, thus the sorptive capacity of foliage was not available to reduce air concentrations. McLachlan and Horstmann (41) have examined the role of forests in removing airborne organic pollutants, and have suggested that SVOCs with log KOA values of >9 will be effectively removed from the atmosphere by forests. PBDEs have reported log KOA values that are typically >9 (30), whereas KOA values of PCBs are generally at least an order of magnitude lower than those of PBDEs for the same number of halogens. For example, the tetra-PBDE PBDE 47 has a reported log KOA value of 10.5 (30), but the tetra-PCB congener PCB 52 has a reported log KOA value of 8.2 (42), a difference of more than 2 orders of magnitude. Thus, we would expect the buffering effect of forests to be more pronounced for PBDEs than for PCBs. Figure 3 appears to support this hypothesis, showing a more dramatic decline in the atmospheric concentrations

of PBDEs than for the PCBs. Wania and McLachlan (16) have modeled this phenomenon, and have suggested that during March and April, prior to bud-burst, the concentrations of SVOCs in air will increase with increasing temperature, and that as new foliage appears in April and May the concentrations will decrease due to an increase in uptake by the vegetation. Thus, we believe that the observed reduction in the concentrations in air of PBDEs and PCBs is closely associated with the emergence of a fresh uncontaminated lipid or wax compartment. The suggestion that PBDEs exhibit an early spring-pulse in air has implications for the manner in which these substances are monitored. Atmospheric samples taken during the early spring in areas that have received snow-fall, or are prone to large diurnal fluctuations in temperature, may give concentration maxima that are atypical of concentrations observed at other times of the year or observed at monitoring sites that are not subject to large temperature fluctuations. This seasonality effect must be considered when evaluating the spatial and temporal distributions of these substances. In addition, there are possible bioaccumulation and toxicity implications of these pulses which should also be considered. The similarities observed in this study between the PCBs and PBDEs cause a concern from the viewpoint of LRAT potential of PBDEs. Whereas ongoing usage/primary emission sources of PCBs to the environment have been reduced substantially, the PBDEs are currently being used in a wide range of consumer products, and as a result, their levels of emission into the environment are increasing. Thus, the quantities of PBDEs transported to remote areas will presumably continue to rise.

Acknowledgments We thank Tom Hutchinson, Eric Sager, and Sheena Symington for their assistance at the James McLean Oliver Ecological Centre; Peter Lafleur for supplying meteorological data recorded at Trent University; Terry Bidleman for providing high-volume samplers and other equipment; three anonymous reviewers for their comments; and NSERC and the consortium of companies that support the Canadian VOL. 36, NO. 7, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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Environmental Modelling Centre for financial support. Lancaster University is grateful to the Department of the Environment, Transport and the Regions (DETR), Air Quality Division, and the Natural Environment Research Council (NERC) Environmental Diagnostics Program for funding.

Supporting Information Available Tables showing the meteorological data collected during the sampling period and summaries of the range and mean of air and leaf-litter concentrations, including blank values, for selected congeners, and a figure showing the location of the sampling site. This material is available free of charge via the Internet at http://pubs.acs.org.

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Received for review July 3, 2001. Revised manuscript received January 14, 2002. Accepted January 15, 2002. ES011105K