Time Trends of Atmospheric PBDEs Inferred from Archived U.K.

These trends were compared to recent modeled estimates of U.K. PBDE emissions. The congener ... Long-Term Trends in PBDEs in Sparrowhawk (Accipiter ni...
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Environ. Sci. Technol. 2005, 39, 2436-2441

Time Trends of Atmospheric PBDEs Inferred from Archived U.K. Herbage ASHRAF HASSANIN,† A. E. JOHNSTON,‡ GARETH O. THOMAS,† AND K E V I N C . J O N E S * ,† Environmental Science Department, Institute of Environmental and Natural Sciences, Lancaster University, Lancaster, LA1 4YQ, United Kingdom, and Rothamsted Experimental Station, Harpenden, Hertfordshire, AL5 2JQ, United Kingdom

Aerial portions of vegetation receive the bulk of their burden of persistent organic pollutants (POPs) from the atmosphere. Vegetation can therefore be a useful indicator of the changing atmospheric burden of POPs. Samples of archived pasture, collected from Rothamsted Experimental Station in the United Kingdom between 1930 and 2004, were analyzed for a range of polybrominated diphenyl ethers (PBDEs). PBDEs could not be routinely detected in the pre1970 samples. Thereafter, the dominant congeners BDE 28, 47, 49, 99, 100, 153, 154, and 183 were frequently detected. The general trend was (a) a rise through the 1970s; (b) a minipeak in the mid-1980s, strongly influenced by one particularly high sample for 1984; (c) values remaining high through the late 1980s/1990s; (d) an indication of a more recent decline for all congeners (except BDE-28), consistent with recent restrictions on PBDE usage in Europe. These trends were compared to recent modeled estimates of U.K. PBDE emissions. The congener profiles of technical mixtures, U.K. air, soil, and pasture were compared and shown to be broadly similar. The implications for environmental release mechanisms are discussed.

Introduction The polybrominated diphenyl ethers (PBDEs) have been widely used as flame retardants in many everyday products, such as furniture, cars, textiles, and other electronic equipment (1-4). They are referred to as “additive” flame retardants, because they are simply blended with the product, as opposed to “reactive” flame retardants that are covalently bound into plastics, for example. This makes them more prone to volatilize into the atmosphere. The global demand for PBDEs has been very substantial; it was estimated to be 70 000 t for the year 1999 (5). The PBDEs have been manufactured and used as three main technical mixtures, the penta- (Pe), octa-, and deca- products (6). The PeBDE technical product, containing Tri-HxBDEs, is currently being considered as a potential/candidate persistent organic pollutants (POPs) under the 1998 United Nations Economic Commission for Europe (UNECE) POPs convention. Some PeBDE congeners appear to have the properties which trigger concerns about their potential long-range atmospheric transport (LRAT), environmental persistence, * Corresponding author phone: 01524 65201; fax: +44 1524 593985; e-mail: [email protected]. † Lancaster University. ‡ Rothamsted Experimental Station. 2436

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tendency to bioaccumulate, and potential toxicity (5, 7). Some may possess the trigger values of physicochemical properties that would categorize them as POPs (8, 9). As a result of risk assessment reports, legislation has been passed to prohibit their marketing and use in the European Union. Europe constitutes an area where extensive use of PeBDE has occurred. Various sampling campaigns have highlighted the occurrence of PeBDE-related congeners in the European atmosphere (10, 11) and people (12-14). A major requirement for understanding the environmental burdens, distributions, and trends of POPs is to have good information on their sources, uses, and estimated emissions (15, 16). Prevedouros et al. (17) recently compiled a European PeBDE emission inventory for the years 1970-2000. Subsequently, they used these estimates in a modeling exercise utilizing the European Variant (EVn) BETR multimedia box model (18) to (a) test their PeBDE emission estimates by investigating the balance between model predictions and ambient measurements; (b) investigate the projected spatial variability in air by using contrasting spatial distribution surrogates; and (c) predict future atmospheric concentration trends using the model in its fully dynamic mode over the period 1970-2010 (19). A key output in their work, however, is the timing and magnitude of the atmospheric emissions profile over time, the emission time trend, which had to be inferred from a series of assumptions. In this study, we present the results from the analysis of archived vegetation (pasture grassland/herbage) samples used to derive a measured time trend and to cross-check the modeled estimates. Aerial portions of vegetation receive the bulk of their POPs burden from the atmosphere (20). Vegetation can therefore be a very useful indicator of the atmospheric burden of POPs during its lifetime/exposure period. In this case, we had access to a unique archived collection from the Rothamsted Experimental Station (see Figure 1), where the long-term Park Grass permanent pasture plots have been sampled, prepared, and stored in the same way every year since the 1840s. Samples from the collection have been analyzed previously to investigate trends of PCBs, PAHs, and PCDD/ Fs (e.g., 21, 22). The pasture samples are particularly useful because they give a direct and dynamic indication of the changes in atmospheric concentrations and hence a direct clue about changing atmospheric burdens (23). In this way, they differ from several other studies of PBDE time trends published in recent years, notably for human tissue (24-26), biota (27, 28), and sediment cores (29, 30). The United Kingdom is a particularly important component of the European inventory for PeBDE. It has been estimated that around 50% of the total European consumption of PeBDE has occurred in the United Kingdom (19); it is therefore not surprising that U.K. ambient air and human tissue concentrations appear to be among the highest in Europe (e.g., 10, 11, 14).

Materials and Methods The Rothamsted Samples. Rothamsted is the oldest active agricultural research station in the world, founded in 1843. It is a semirural location, 42 km north of central London in southeast England (see Figure 1). All the samples were taken from the Park Grass permanent grassland experiment. Park Grass is cut twice a year for hay in June and September. Only the June samples were analyzed for this study. Samples are air-dried after harvest, chopped into short lengths, and then transferred to sealed glass jars or metal tins in the Rothamsted archive for storage. Hence, the loss of PBDEs during storage is not believed to be important. The key issue concerning 10.1021/es0486162 CCC: $30.25

 2005 American Chemical Society Published on Web 03/01/2005

FIGURE 1. Location map of Rothamsted Experimental Station. sample integrity and their ability to reflect ambient air is the potential for contamination. This has been discussed in some detail previously in relation to studies on other POPs (3133). One issue is the possibility for postcollection contamination during storage, via dust. We are confident that this does not occur, because the storage vessels are sealed and all traces of dust are removed at the time of subsampling (see refs 31-33 for discussion). A second issue concerns the air-drying step (34, 35), which has always taken place in one of the buildings at the station. For this study, subsamples were taken from the archive for the years 1903, 1930, 1940, 1950, 1960, and 1961-2004, 49 samples in total. Sample Extraction; Cleanup and Analysis. Approximately 10 g of powdered grass was mixed with Na2SO4 (1:3), transferred to a pre-extracted glass thimble, and spiked with a recovery standard containing 13C12 PCBs 138, 153, 180, and 209. The samples were extracted for 20 h using toluene. The toluene was evaporated and the extracts were transferred with hexane to clean 20-mL vials with 1 mL of sulfuric acid (H2SO4) and left overnight for lipid and wax digestion. The extract was then further cleaned with multilayer silica columns (30-mm i.d.) filled from the bottom with 1 cm Na2SO4, 1.5 g activated silica, 3 g silica/NaOH (1 M), 1.5 g activated silica, 6 g silica/H2SO4 44% (w/w), and 1.5 g silica topped with 1 cm Na2SO4 eluted with 200 mL hexane. The concentrated samples were applied twice to a 6 g Biobead (SX-3) column to remove any remaining lipids. The sample was eluted with 50 mL hexane/DCM (1:1), the first 16 mL was discarded, and the following 30-mL fraction was collected in an amber vial. The sample was then concentrated and applied for a second time to the same column after flushing with another 50 mL hexane/DCM (1:1). The first 16 mL was discarded and the following 30-mL fraction was collected in the same amber vial. These fractions were then used for PBDEs determination. Twenty-five microliters of dodecane containing internal standards was added to the sample volume, which was then reduced in volume to 25 µL for analysis. The samples were analyzed for PBDEs by GC-MS on a Fisons MD800 operated with a negative chemical ionization source in selected ion monitoring mode, using ammonia as the reagent gas. The following PBDE congeners were in the standard mix (17, 28, 32, 35, 37, 47, 49, 66, 71, 75, 77, 85, 99, 100, 119, 138, 153, 154, 166, 181, 183, 190). The following congeners were detected in many of the samples: tetra-BDE

47; penta-BDE 99, 100; hexa-BDE 153, 154. The sum of these five congeners is referred to subsequently as Σ5PBDE. Congeners 17, 28, 32, 35, 37, 49, 66, 71, 75, 77, 85, 119, 138, and 183 were also detected, though less frequently. The sum of these congeners and Σ5PBDE are referred to as ΣALLPBDE. BDE 166, 181, and 190 were not detected in any of the samples. Most of these congeners are present in the penta-BDE mix (6). BDE-183 is a marker of the octa-BDE product (6) and so will have different sources to the environment. It is therefore considered separately in the discussions below. QA/QC. Recoveries were monitored for all the samples using a mixture containing 13C12 PCBs 138, 153, 180, and 209 that was added to the sample prior to extraction. Average recoveries were 93% for 13C12 PCB 138, 74% for 13C12 PCB 153, 110% for 13C12 PCB 180, and 79% for 13C12 PCB 209. Blanks (extraction of Na2SO4 in a thimble) were included at a rate of one blank for every five samples. All results were blankcorrected. The blank recoveries were 94% for 13C12 PCB 138, 72% for 13C12 PCB 153, 112% for 13C12 PCB 180, and 84% for 13 C12 PCB 209. The practical detection limit for all PBDE congeners ranged between 6 and 50 pg/g. The reproducibility of the method was good; replicate extraction (n ) 3) of the 1989 sample gave an average RSD of 15% (range 1-37% for different congeners).

Results and Discussion General Comments on the PBDE Trends and Concentrations. All data are expressed on a dry weight (DW) basis. The general trend was of nondetectable levels of PBDEs in the early samples, increasing ΣALLPBDE concentrations in the 1970s, highest levels in the 1980/90s, and recent declines. The ΣALLPBDE values in grass varied substantially, by a factor of ∼120, between 10 and 1200 pg/g. The following congeners were detected in most samples: BDE 28, 47, 49, 99, 100, 153, 154, and 183. BDE-47 and -99 generally dominated the mixture, with BDE-28, -35, -100, and -153 generally contributing >25% of the total. Other congeners BDE 17, 32, 35, 66, 71, 75, 77, 85, 119, and 138 were also detected, although less frequently. Comments on the Introduction of PBDEs into the Environment and the Early Samples. The early samples were taken from the environment before the commercial production and use of PBDEs began. Some detectable values were obtained in some of the 1960s samples, but these were erratic and do not represent proof of PBDE occurrence in the environment in the early sample years. Nylund et al. (29) reported BDE-47 and -99 in sediment layers dating from the 1930s to the late 1950s, although there is obviously the potential for deeper sediment deposits in cores to become mixed with fresher, contaminated material from above. PBDEs were not measurable in any of the 420 archived serum samples collected in the 1960s from the San Francisco Bay area and analyzed recently (13). Natural production of some POPs has been reported; indeed, different species of marine sponge have been reported to produce brominated compounds similar to PBDEs (36-38). Some ambiguities appear in the literature, as to when industrial production of the PBDEs began. Prevedouros et al. (17, 19) assumed it to be the early 1970s in Europe for their consumption and emission calculations, on the basis of available production and use information. However, Kierkegaard et al. (28) reported low levels of BDE-47, -99, -153, and -154 in archived Swedish aquatic biota samples from 1967 to 1969. Similarly, trace levels were present in Swedish human breast milk samples collected in 1972 (39), while Kajiwara et al. (27) reported low levels of PBDEs in the earliest (from 1972) northern fur seal samples they analyzed from the Pacific coast of Japan. It is clear from some of the early literature that brominated flame retardants were VOL. 39, NO. 8, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 2. Time trend for the measured ΣALLPBDE (a) and BDE-47 (b) in archived pasture versus the modeled trend of European atmospheric emissions (ref 19). attracting attention from the analytical chemistry and toxicology community by the 1970s (e.g., 40, 41). Figure 2a presents the measured ΣALLPBDE trend in the grass samples and the projected atmospheric emissions calculated by Prevedouros et al. (17, 19) on the basis of estimates derived for BDE-47 (see below). The routine detection of BDE congeners in the archived grass samples begins in the 1970s and generally rises thereafter. However, while the modeled emissions of Prevedouros et al. (17, 19) were projected to rise steadily, peak quite sharply, and then decline, the measured data indicate a longer, flatter emission profile (see later discussion). Comments on the General Trends. The analytical data are quite variable, year to year (see Figures 2 and 3) (see also ref 42). This is not surprising because several factors combine to influence herbage concentrations. These include air concentrations (i.e., the variable of interest) and an array of factors which potentially affect air-plant transfer (e.g., particle trapping and retention by vegetation, temperature, rainfall, etc.) (see refs 20-23). Figure 3 shows data for five selected congeners: BDE-28, -47, -99, -153, and -183. The data show (a) a rise through the 1970s; (b) a minipeak in the mid-1980s, strongly influenced by one particularly high sample for 1984; (c) values remaining high through the late 1980s/1990s; and (d) an indication of a more recent decline for all congeners except BDE-28. These broad trends agree very closely with those reported by Kierkegaard et al. (28) for Swedish aquatic biota; they also observed high levels in the 1980s, for example. Trends in northern fur seals from the Pacific coast of Japan peaked later in the early/mid-1990s (27). 2438

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Recent Declines. As noted above, the peak year in the 1990s grass samples occurred in 1999. Evaluation of the ΣALLPBDE trends (minus BDE-28, excluded for reasons mentioned below) indicates a decline of 75% between 1999 and 2004. European data generally seem to have peaked and to be slowly declining (this study; 28, 39). This is in contrast to the situation in North America, where recent reports highlight continuing and substantial increases in human milk (43), marine mammals (44, 45), and birds eggs (46), for example. This is consistent with known differences in use trends between the continents. In Europe, PeBDEs have been restricted and banned for several years now, whereas voluntary restrictions are only now beginning to come into force in North America (1, 17, 19, 47). Congener-Specific Observations. As just noted, one exception to the broad trends noted above is for BDE-28 (Figure 3a). While other congeners have fallen in recent years (Figures 2, 3B-D), BDE-28 alone increases in concentration. It makes quite a substantial contribution to the ΣALLPBDE, which masks the recent declines of the other congeners (see Figures 2 and 3). Possible explanations presumably include additional recent sources of BDE-28 in other brominated products and possible formation of BDE-28 from other compounds in the environment. The modeled emission trend shown in Figure 2b was originally derived for BDE-47 in the study by Prevedouros et al. (17). It is therefore also shown superimposed on the BDE-47 measurements plot in Figure 2b, where it generally gives a better fit than was seen for the ΣALLPBDE data in Figure 2a. BDE-99 and -153, other major constituents of the

FIGURE 3. Time trends for the congeners: (a) BDE-28; (b) BDE-47; (c) BDE-99; (d) BDE-153; (e) BDE-183.

FIGURE 4. Comparison of the congener profile in the technical PeBDE mixture (ref 6), summertime U.K. air (ref 11), pasture grass (this study), and soil (ref 48). PeDBE technical mixtures (6), follow the BDE-47 trend, and all indicate a post-1990s decline. It is also appropriate to consider BDE-183, since it is a marker for the octa-BDE mix, as opposed to the Pe-BDE mix (6). The trend for BDE-183, shown in Figure 3e, suggests that it is introduced into the environment slightly later than the

other congeners (i.e., around 1980). Like the PeBDE congeners, it also shows a long, flat trend, rather than a sharp pulsed input. This is discussed further below. Comparing the PBDE Patterns in Grass, Air, and Soil. While several papers report the concentrations of PBDEs in air, almost no data have appeared for vegetation or soil. Given VOL. 39, NO. 8, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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the importance of air-vegetation and air-soil exchanges in controlling POPs fate and behavior in background environments, the congener composition of these media are therefore compared here. Figure 4 summarizes the patterns of BDE47, 99, 100, 153, and 154 in the Rothamsted grass samples (averaged for the years 1998-2004) to that in southern U.K. summertime air from 2000 (11), U.K. grassland soils sampled in the late 1990s (48), and the technical Bromkal 70-5DE mixture, reported by Sjo¨din et al. (6). It is clear that the congener profile of pasture grass broadly reflects that of air which, as noted before (11, 48), broadly reflects that of the technical PeBDE mix. This indicates that there is very little difference in the overall source-pasture transfer efficiency of the different congeners, despite substantial differences in those physicochemical properties (i.e., vapor pressure, partition coefficients) which are assumed to affect their emissions and environmental fate. This is a very important observation and is discussed further below. The soil congener profile is somewhat different, being relatively enriched in BDE-99, BDE-153, and BDE-154. We speculate that this may be because (a) BDE-47 is degraded relatively rapidly in soils (enhancing the relative proportion of BDE-99) or (b) additional wintertime, combustion-related sources of BDE-99, BDE-153, and BDE-154 (11, 49) may also accumulate in soils and be reflected in the soil pattern (48), while this is not seen in air and pasture sampled in the summer. Implications for Sources. As noted in the Introduction, the chief value of this study is in providing data from which inferences about sources can be made. We offer the following observations: 1. The trends in air concentrations, inferred from the pasture samples, suggest a flatter emission profile than predicted by Prevedouros et al. (17, 19). Their modeling study suggested that the timing and magnitude of the emission peak is a strong function of the assumptions made about the lifetime of treated PBDE products; shorter product lifetimes, or greater emissions from the product during its lifetime, will tend to flatten the projected emission curve (17). 2. The similarities in congener pattern between technical mixtures, air, and pasture is important. The modeled atmospheric emissions were assumed to be dominated by volatilization from products. Furthermore, it was assumed that these are proportionately greater for the lower molecular weight BDEs (17, 19). Specifically, volatilization from treated polyurethane foams (PUF) and textiles were assumed to be the dominant emission mechanism. However, taking the observations on congener patterns together with the observations under 1, it may be that one or both of these assumptions are incorrect. One alternative scenario may be that physical weathering of PUF as it ages, with the associated crumbling of the foam (50, 51), generates a fine light dust which ultimately gets dispersed into the atmosphere. This emission and transport mechanism would presumably preserve the congener pattern from the foam to the atmosphere much more effectively than vapor pressure or octanol: air partition coefficient controlled volatilization from PUF (17). Alternatively, modifications in the gas:particle distribution of PBDEs following release from source areas may have an important role in their subsequent fate, once reaching rural terrestrial systems. A further possibility is that emissions are indeed congener-specific and favor the lighter ones which volatilize preferentially from the foam, but they are also preferentially degraded in the atmosphere (52-54). 3. Temporal trends will presumably differ between location and media. Other trend studies, on sediment cores and biota in Europe, tend to suggest a sharper, pulsed emission than this study. However, many of these studies provide data on trends in samples which are more reflective of discharges/trends in aquatic systems (e.g., 27, 28, 30, 4446). 2440

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This study raises some important questions about the timing, magnitude, and mode of primary atmospheric emissions of PeBDE congeners to the U.K. terrestrial environment. Until these are resolved, there will be uncertainties hampering predictions about likely future trends in environmental levels and exposures. Further studies are required to elucidate these issues.

Acknowledgments We are grateful to the Egyptian Ministry of Higher Education for doctorate funding for A.H.

Supporting Information Available PBDE data (pg/g dwt.) in the full sample set. This material is available free of charge via the Internet at http:// pubs.acs.org.

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Received for review September 5, 2004. Revised manuscript received January 14, 2005. Accepted January 24, 2005. ES0486162

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