Concentrations of Polychlorinated Biphenyls in Indoor Air and

Our principal objectives were (1) to significantly augment the worldwide ... Our purification method was based on that developed for PCB and PBDE anal...
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Environ. Sci. Technol. 2006, 40, 4633-4638

Concentrations of Polychlorinated Biphenyls in Indoor Air and Polybrominated Diphenyl Ethers in Indoor Air and Dust in Birmingham, United Kingdom: Implications for Human Exposure STUART HARRAD,* SADEGH HAZRATI, AND CATALINA IBARRA Division of Environmental Health and Risk Management, Public Health Building, School of Geography, Earth, and Environmental Sciences, University of Birmingham, Birmingham, B15 2TT, United Kingdom

Polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) were measured in air (using PUF disk passive samplers) in 31 homes, 33 offices, 25 cars, and 3 public microenvironments. Average concentrations of ΣBDE (273 pg m-3) and ΣPCB (8920 pg m-3) were an order of magnitude higher than those previously reported for outdoor air. Cars were the most contaminated microenvironment for ΣBDE (average ) 709 pg m-3), but the least for ΣPCB (average ) 1391 pg m-3). Comparison with data from a previous spatially consistent study, revealed no significant decline in concentrations of ΣPCB in indoor air since 199798. Concentrations in indoor dust from 8 homes were on average 215.2 ng ΣBDE g-1, slightly higher than other European dust samples, but twenty times lower than Canadian samples. Inhalation makes an important contribution (between 4.2 and 63% for adults) to overall UK exposure to ΣPCB. For ΣBDE, dust ingestion makes a significant buts in contrast to Canadasa not overwhelming contribution (up to 37% for adults, and 69% for toddlers). Comparison of UK and Canadian estimates of absolute exposure to ΣBDE suggest that differences in dust contamination are the likely cause of higher PBDE body burdens in North Americans compared to Europeans.

Introduction Polychlorinated biphenyls (PCBs) have found widespread use in a diverse range of applications, with around 1.2 million tonnes produced worldwide (1). Of this, approximately 67 000 and 40 000 t were produced and used, respectively, in the UK (1). Owing to concerns about their adverse effects on humans and wildlife, their productionsbut not their uses ceased in the UK and throughout most of the industrialized world in the late 1970s. Although UK human exposure to PCBs via the diet has fallen in recent years in response to the cessation of their production (2, 3), human health concerns remain: currently a substantial proportion of UK schoolchildren and toddlers are exposed to dioxins and dioxin-like * Corresponding author e-mail: [email protected]; tel: +44 121 414 7298; fax: +44 121 414 3078. 10.1021/es0609147 CCC: $33.50 Published on Web 06/29/2006

 2006 American Chemical Society

PCBs via the diet at levels that exceed the UK government’s recommended tolerable daily intake (3). Although the majority of nonoccupational exposure to PCBs has been widely considered to occur via the diet (4), there have been increasing indications that indoor air remains contaminated by PCBs remaining in use in applications such as permanent elastic sealants and acoustic ceiling tiles (5-7). As a result, inhalation of indoor air could constitute a significant exposure pathway; a fact compounded by the downward trend in dietary exposure. Polybrominated diphenyl ethers (PBDEs) have found wide use as flame retardants. In recent years, production and use of PBDEs has been in the guise of three formulations: penta (consisting primarily of BDEs 47 and 99, 37% each, alongside smaller amounts of other tetra-, penta-, and hexa-BDEs), octa (a mixture of hexa- (10-12%), hepta- (44-46%), octa(33-35%), and nona- (10-11%)), and deca (98% decabromodiphenyl ethersBDE 209sand 2% various nona-BDEs) (8, 9). Worldwide, PBDE production is dominated by the deca commercial formulation, with global demand in 2001 an estimated 56 100 t (10). This is similar to the 1999 estimate of 54 800 t (11). By comparison, 2001 global demand for the penta-product was 7500 t (10), down slightly from 8500 t in 1999 (11). Production and use of commercial PBDE formulations in Europe was considerably less than that in North America, for example, in 2001, 7100 t of penta-product was used in North America, compared to just 150 t in Europe (10). The uses for these commercial formulations are myriad: the penta-product was employed principally to flame-retard polyurethane foams in carpet underlay, vehicle interiors, furniture, and bedding; the octa-formulation was used to flame-retard thermoplastics such as high-impact polystyrene, and the deca-product was used principally in plastic housings for electrical goods such as TVs and computers, as well as in textiles (8). As a result of concerns surrounding these contaminants owing to their presence in the diet and indoor air and dust (12-14), and human tissues (15), coupled with evidence relating to their potential adverse effects on human health (9, 16), several jurisdictions have banned the marketing and use of penta- and octa-BDEs. Furthermore, the main U.S. producer and the U.S. EPA have reached a voluntary agreement to discontinue production of the penta- and octa-BDE mixtures. Despite this, there remain strong concerns that the existing (largely indoor) reservoir of PBDEs (and PCBs) associated with treated goods represents a substantial source of current and future exposure to these compounds, both via direct inhalation and ingestion of contaminated indoor air and dust, and in time via dietary exposure following their emission, transport, and incorporation into the diet (17). This study reports concentrations of a number of individual PCB and PBDE congeners in indoor air (sampled using PUF disk passive air samplers) taken from 31 homes, 33 offices, 25 cars, and 3 public microenvironments within the West Midlands, the UK’s second largest conurbation (population 2.5 million). These data are combined and compared with existing data on UK dietary exposure to PCBs to provide a preliminary indication of the relative significance of inhalation and diet to overall human exposure to PCBs for UK adults and toddlers. Further, given recent indications that, for Canadians, ingestion of indoor dust may constitute the most important exposure pathway to PBDEs for many individuals (13), concentrations of PBDEs in dust samples taken from 8 West Midlands homes are reported, and used in conjunction with estimates of inhalation and dietary intakes to provide a preliminary indication of the relative VOL. 40, NO. 15, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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significance of dust ingestion, diet, and inhalation as pathways of exposure to PBDEs for UK adults and toddlers. This study focused on BDEs 28, 47, 49, 66, 85, 99, 100, 153, and 154. These congeners were selected on the basis that they are the principal congeners monitored in previous comparable studies (18-20). Although decabromodiphenyl ether is being increasingly reported, it was not included in this study owing to the difficulties in achieving its reliable determination at the outset of the study (21). The PCB congeners monitored are in line with those reported previously by our research group (5, 22, 23). Our principal objectives were (1) to significantly augment the worldwide database on concentrations of PBDEs and PCBs in indoor air; (2) to evaluate the significance of inhalation as a pathway of exposure to PCBs of UK adults and toddlers, both relative to dietary exposure and to previous estimates of such exposure; (3) to evaluate the temporal trend in concentrations of PCBs in indoor air from the West Midlands; and (4) to evaluate the significance of inhalation and dust ingestion as a pathway of human exposure to PBDEs of UK adults and toddlers, both relative to other exposure pathways and to similar evaluations in Canada.

Experimental Section Sampling Strategy. Indoor air samples were collected between September 2003 and November 2005 in a total of 92 microenvironments within the West Midlands conurbation. At least one sample was taken in each microenvironment; where more than one sample was taken in a given microenvironment, we have used the average concentration. Samples were all taken under normal room use conditions, thus reflecting actual human exposures as far as possible. From previous knowledge of activity patterns of West Midlands residents (24), we selected the following microenvironment categories for study (public transport was considered unsuitable for siting the passive air samplers employed): homes (both living rooms and bedrooms; n ) 31), offices (n ) 33), cars (n ) 25), and public microenvironments (n ) 3; a coffee shop, a supermarket, and a post office). The cars represented a wide variety of models and manufacturers: 5 Nissans, 3 Toyotas, 3 Renaults, 3 Fiats, 2 Woleks, and 9 others. The average age of the vehicles studied was 10 ( 5 years ranging from 1 to 21 years. Samples of dust were also taken from 8 homes, selected at random from the microenvironments monitored for indoor air contamination. Passive Air Sampling. Passive air samplers (i.e., PUF disks) were employed to provide a time-integrated sample over each 28 day sampling period. These have been used successfully in other studies of POPs in indoor air (18, 19). Samplers (each comprising one shelter, each fitted with one PUF disk) were deployed in homes, offices, and public microenvironments at a height of approximately 1 m above the floor, supported by a customized stainless steel cradle that permitted air flow to all sides of the sampler. For practical and safety reasons, samplers were deployed in the trunks of cars rather than in the passenger/driver compartments, again supported by a secured stainless steel cradle. While there may be intra-vehicle variations in concentrations, we have assumed those detected in the trunk to reflect that inhaled by the driver and passengers. Each PUF disk measured 14 cm in diameter and 1.2 cm in thickness, giving a surface area of 360 cm2, and density of 0.01685 g cm-3. Disks were sheltered by two different size stainless steel housings (18 cm, 1 L bottom housing and 23 cm, 2 L top housing). Before deployment, disks were washed thoroughly in tap and distilled water sequentially to remove loose material, then extracted in hexane using a Soxhlet apparatus for 48 h to remove target or interfering compounds. Following extraction, disks were desiccated to remove solvent, spiked with 4634

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known quantities of PCBs 19 and 147 as QA/QC standards to provide a measure of contaminant loss during sampling, and stored in pre-cleaned foil in airtight solvent-cleaned glass jars. On deployment, disks were removed from the jars on site and transferred into the shelters. At the end of each sampling period, disks were removed from shelters and stored in solvent-cleaned aluminum foil in airtight glass jars at 4 °C until extraction. Conversion of contaminant masses per sample into concentrations in air requires knowledge of the air sampling rate of the PUF disk samplers and their deployment time. We used sampling rates reported previously from a specific calibration exercise conducted using our sampler configuration in an office (25). To summarize, individual congenerbased sampling rates ranging between 1.1 and 1.9 m3 day-1 were used for PBDEs; while for PCBs, homologue groupbased sampling rates in the range 0.70-1.27 m3 day-1 were employed. This is because sampling rates for individual PCB congeners of the same homologue group were very similar. Indoor Dust Sampling. Dust samples were collected from 8 homes in the autumn of 2005. In each home, inhabitants were asked to provide the contents of their vacuum cleaner bags. Following collection, each sample was transported to the laboratory, passed through a 500 µm mesh size sieve, weighed, and stored in a clean glass container at (-18 °C) until analysis. Before removal of a 1 g aliquot for analysis, each sample was shaken thoroughly. Analytical Protocols. Our purification method was based on that developed for PCB and PBDE analysis (12, 26) with minor modifications. In summary, air samples were treated with appropriate quantities of PCB and PBDE internal/ surrogate standards (specifically PCBs 34, 62, 119, 131, and 173, and 13C PBDEs 28, 47, 99, and 153) before extraction in a pre-cleaned Soxhlet apparatus for 8 h using HPLC grade hexane. For dust samples, accurately weighed 1 g aliquots of each sample were extracted using hexane by an accelerated solvent extraction (ASE) system (ASE 300, Dionex), using a 66 mL cell filled from the bottom with Florisil (1.5 g), sample, and Hydromatrix (Varian Inc.). Extraction conditions were as follows: temperature 150 °C, pressure 1500 psi, heat time 7 min, static time 5 min, flush volume 50%, purge time 100 s, static cycles 1. Following extraction, crude extracts were concentrated to approximately 2 mL, treated with 2 mL of concentrated sulfuric acid, subjected to liquid/liquid back extraction using dimethyl sulfoxide, prior to elution through a column containing 1 g of Florisil (Aldrich Chemicals; 60-120 mesh, pesticide grade) topped with 1 g of anhydrous sodium sulfate with 20 mL of hexane. The eluate was reduced to incipient dryness before addition of 20 µL of nonane containing 10 ng each of PCBs 29 and 129 as recovery determination standards (RDSs), against which recoveries of internal/surrogate standards may be measured. PCB and PBDE analyses were conducted on a Fisons’ MD-800 GC/MS system fitted with a 60 m VF5 MS column (0.25 mm id, 0.25 µm film thickness). Both injector and interface temperatures were 280 °C. The oven temperature program for PCBs was as follows: 140 °C for 2 min, 5 °C/min to 215 °C and held for 5 min, then 2 °C/min to 280 °C and held for 15 min. PBDE separation was achieved by a temperature program of 140 °C with 5 °C/min ramp to 200 °C and 2 °C/min to 300 °C held for 10 min. The mass spectrometer was operated in EI+ SIM mode; with monitored m/z values as reported previously (12, 26). To ensure accurate and precise measurement, peaks were only accepted if the following criteria were met: (1) signalto-noise ratios for the least abundant ion exceeded 3:1; (2) peaks eluted within 5 s of standards run in the same batch as the samples; (3) isotope ratios for peaks were within 20% of those obtained for standards run in the same batch as the samples.

TABLE 1. Summary of Concentrations (pg m-3) of ΣPCBs and ΣBDEs in Air Samples from Different Indoor Microenvironment Categories in This and Selected Other Studies statistical parameter/ study location (reference)

all MEs all MEs (excluding (excluding cars) cars) ΣBDE ΣPCB

minimum, this study average, this study σn-1, this study 5th percentile, this study median, this study 95th percentile, this study maximum, this study Birmingham, UKa W. Midlands UK, 1996-1998a Switzerland, 2001b

4 110 208 6 47 358 1416

W. Midlands, UK, 2001-2003c Ottawa 2002-2003d Kuwait, 2004e

762

487 10725 20714 613 3538 43753 101762

offices ΣBDE 10 166 275 14 71 638 1416

offices ΣPCB

homes ΣBDE

816 4 18149 52 26856 61 1131 5 5901 24 87706 182 101762 245

public public homes MEs MEs cars cars ΣPCB ΣBDE ΣPCB ΣBDE ΣPCB 487 2823 2568 594 1802 8899 9764

29 112 72 41 144 160 162

outdoor ΣBDE

outdoor ΣPCB

1078 11 392 30732 709 1391 44213 1866 1206 1927 12 483 9570 41 929 74351 4233 2575 81548 8184 6018 252

9000 790000 (410000)

14512 2778 32.7 (8.6)

3016 525 260 (100) 15.2 (8.2)

21 2.2 (2.6)

a

Using high-volume air samplers, sum of same range of PCB congeners monitored in the current study (Birmingham, average, ref 23; W. Midlands, average, ref 5). b Obtained using low-volume active air samplers, ΣPCB ) 5 × (Σ congeners 28, 52, 101, 138, 153, and 180) (average (median in parentheses), ref 7). c Obtained using high-volume air samplers, sum of PBDEs 47, 99, 100, 153, and 154 (average, ref 12). d Obtained using PUF disk passive samplers, sum of PBDEs 17, 28, 47, 66, 71, 85, 99, 100, 153, and 154 (average (median in parentheses), ref 19). e Obtained using PUF disk passive samplers, sum of PBDEs 28, 47, 99, 100, 153, 154, and 183 (average (median in parentheses), ref 18).

For air samples, field blanks consisting of a PUF disk (treated in identical fashion to those used for sampling, except that no air was aspirated through them) (n ) 19) and method blanks (i.e., same as field blanks but PUF disks were not transported to/from sampling site) (n ) 4) were found to contain concentrations of target PBDEs and PCBs no greater than 4% of the concentrations found in the corresponding samples. Our data are thus not corrected for blank concentrations. Method blanks (n ) 2) for the dust analysis procedures consisting of sodium sulfate were found to be similarly low in contamination. Average recoveries of internal standards for all samples ranged from 45% (13C12-BDE-153) to 67% (13C12-BDE-47), while those of the QA/QC standards (PCBs 19 and 147) added to PUF disks prior to sampling to provide an indication of measure of contaminant loss during sampling and analysis combined, were 95 and 75% respectively. Air sample concentrations were not corrected for such losses. The repeatability of our passive sampling and analytical procedures combined was evaluated by simultaneously deploying 4 passive samplers at the same location. The low relative standard deviations observed for concentrations of the target PBDE congeners (average ) 3.8%; range 0.9-6.0%) demonstrate good repeatability for our sampling and analytical method. Method detection limits for individual BDEs were typically 0.1 pg m-3 and 0.03 ng g-1 for air and dust respectively, while those for individual PCBs were typically 0.1 pg m-3 for air. The accuracy of our methods is indicated by our satisfactory performance in the 2002 BSEF/ QUASIMEME interlaboratory comparison on brominated flame retardants (12), and by our previous studies on PCBs in air (5, 22, 23).

Results and Discussion PBDE Concentrations in Air from Different Categories of Indoor Microenvironments. The average concentration of ΣBDE for all indoor microenvironments studied (i.e., cars, homes, offices, and public microenvironments combined) is 273 pg m-3 with a median of 46 pg m-3. Table 1 summarizes the concentrations of ΣPBDE in air samples taken in this study from each of the four microenvironment categories studied, and compares them with those recorded in other relevant studies. Compared to these other studies, concentrations in this study are noticeably higher than those in outdoor air in the West Midlands (12) and indoor air in Kuwait (18), but lower than those in indoor air in Ottawa (19), and in a previous, preliminary study in the West Midlands (12).

We believe that the lower concentrations in this study are unlikely to be indicative of a temporal trend, and that they arise as a result of the comparatively limited numbers of samples taken in the earlier study. With respect to congener pattern, the major contributors to ΣBDE are congeners 47 and 99. It has been reported that preferential volatile emissions of the lower brominated 47 cf. 99 from household items favor higher ratios in air compared to those detected in the treated material itself (27), which for the penta-BDE DE-71 and Bromkal 70-5DE formulations employed in the UK are ∼0.7 and ∼1.0, respectively (28, 29). Our data are consistent with this hypothesis, as the median ratio of congeners 47:99 in all indoor air samples analyzed in this study is 1.2. As indicated by our previous study of PBDEs in indoor air in the West Midlands (12), concentrations in offices in this study exceed those in homes (p < 0.05 Mann-Whitney U-test). Overall, concentrations in cars are the highest of the four microenvironment categories studied, although the average concentration is heavily influenced by the contribution of three cars where concentrations are 2.57, 4.65, and 8.18 ng m-3 respectively. Although to our knowledge the data given here are the first report of concentrations of PBDEs (and PCBs) in air sampled in cars, by demonstrating that car interiors are appreciably contaminated with PBDEs, they are consistent with the only other available information on PBDE contamination of windshield film and dust sampled from the interior of U.S. cars (30). On the basis of evidence to date, further monitoring of PBDE contamination in cars is warranted. PCB Concentrations in Air from Different Categories of Indoor Microenvironments. The average concentration of ΣPCB for all indoor microenvironments studied (i.e., cars, homes, offices, and public microenvironments combined) is 8.92 ng m-3 with a median of 2.53 ng m-3. Table 1 summarizes the concentrations of ΣPCB in air samples taken in this study, and compares them to those recorded in other relevant studies. Compared to these other studies, concentrations in this study are well above those detected in outdoor air in Birmingham (22, 23), similar to those previously reported in indoor air in the West Midlands (5), but substantially lower than those recently found in a survey of buildings suspected to contain PCB-treated joint sealants in Switzerland (7). The congener pattern is dominated by the tri- and tetrachlorinated homologues, which for all samples in this study constitute a median of 91% of ΣPCB, suggesting VOL. 40, NO. 15, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 2. Summary of Estimates of Exposure (ng day-1) of UK Adults and Toddlers to ΣPCB via Inhalation and Diet, and Relative Significance (%) of Each Pathway adult

toddler (6-24 months) d-1)

5th percentile air food Σ exposure air food

15 340 355 4.2 95.8

median 60 340 400 15.0 85.0

average

intake (ng 95th percentile

150 340 490 30.6 69.4

586 340 926

5th percentile 2.8 194 197

% contribution 63.3 36.7

1.4 98.6

median 11 194 205 5.4 94.6

average 28 194 222 12.6 87.4

95th percentile 111 194 305 36.4 63.6

TABLE 3. Concentrations (ng g-1 DW) of ΣBDE in Dust Samples in This and Selected Other Studies statistical parameter, study location minimum, this study average, this study σn-1, this study 5th percentile, this study median, this study 95th percentile, this study maximum, this study

a

concentration 16.2 215.2 243.6 17.8 87.1 586.7 625.4

statistical parameter, study location (reference)

concentration

minimum, Ottawaa(20) maximum, Ottawaa(20) average, Ottawaa(20) minimum, Spainb(33) maximum, Spainb(33) average, Spainb(33) minimum, Belgiumb(33) maximum, Belgiumb(33) average, Belgiumb(33)

64 170000 4500 2.9 380.2 36.6 6.2 384.8 58.7

Sum of PBDEs 17, 28, 47, 66, 85, 99, 100, 138, 153, 154, 183, and 190. b Sum of mono- through hepta-BDEs.

that a lower chlorinated commercial formulation such as Aroclor 1242 is the predominant source in the indoor environments studied here. Statistical comparison (using a Mann-Whitney U-test) of concentrations of ΣPCB in offices and homes in this study confirm offices are more contaminated than homes (p < 0.001). In contrast to PBDEs, the concentrations of PCBs detected in cars are the lowest of all four microenvironment categories studied, implying that sources of PCBs in UK cars are not significant. Inhalation exposure to PCBs. To evaluate the likely magnitude of external exposure via inhalation to PCBs to the inhabitants of the West Midlands, we have assumed 100% absorption of intake. We have used estimates of adult human time activity patterns derived from unpublished data produced by an ongoing study in our research group of 70 individuals living in the West Midlands covering both summer and winter (31). This shows the average percentage of time spent in homes, offices, public microenvironments, transportation microenvironments (including cars), and outdoors to 63.8, 22.3, 5.1, 4.1, and 4.5%, respectively. We have then estimated various plausible inhalation exposure scenarios, based on a daily air inhalation rate of 20 m3 day-1, and using 5th percentile, median, average, and 95th percentile concentrations in each microenvironment category, including outdoors. The range of exposure estimates thus derived is only an indication of the likely range within the population, as this study has only sampled a very small proportion of individual microenvironments, and exposure of individuals will also be strongly influenced by their specific time activity pattern and the level of contamination in the specific microenvironments they frequent. In the absence of information on the time activity patterns of toddlers (defined as 6-24 months), exposure to this population group has been estimated in a fashion similar to that for adults, but using a daily air inhalation rate of 3.8 m3 day-1 (20), and assuming the average percentage of time spent in homes, offices, public microenvironments, transportation microenvironments (including cars), and outdoors to be 86.1, 0, 5.1, 4.1, and 4.5%, respectively. Table 2 summarizes these estimates of the exposure of both adults and toddlers to PCBs via inhalation, both in 4636

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absolute terms and relative to dietary exposure. Dietary exposure for adults has been assumed to be 340 ng ΣPCB d-1 for adults (32), and that of toddlers is assumed to be 57% of that of adults (20). We have not considered the impact on exposure that would arise if the toddler were breast-fed. From Table 2, it appears that inhalation can make an importants albeit not overwhelmingscontribution to the sum of inhalation and dietary exposure of adults (up to 63%) and to a lesser extent of toddlers (up to 36%). The figures for both absolute adult exposure and the relative significance compared to diet are in striking agreement with the figures (average daily inhalation exposure ) 110 ng; representing 24% of the sum of inhalation and dietary exposure) derived from a similar study conducted in the West Midlands in 19971998 (5). Temporal Trends in Concentrations of PCBs in Indoor Air. We have previously reported on temporal trends in concentrations of PCBs in outdoor air (23). To our knowledge however, similar analysis has not been previously conducted for indoor air, either in Birmingham or elsewhere. We therefore compared concentrations from a previous study conducted by our group, which reported concentrations of PCBs in indoor air samples from 5 offices, 2 laboratories, and 7 homes in the West Midlands between December 1996 and March 1998 (5). Although the numbers and exact locations of microenvironments are not directly comparable, a statistical comparison of concentrations from the current study (excluding cars) and the previous study reveals no statistically significant difference (Mann-Whitney U-test; p > 0.1) in ΣPCB concentrations in the two studies. While further, more detailed studies of temporal trends are required, this study suggests there has been no discernible reduction in the contamination of indoor air in the West Midlands of the UK since 1996-98. This contrasts with the marked downward temporal trend in PCB contamination of the UK diet (2, 3). Concentrations of PBDEs in Indoor Dust. Table 3 summarizes concentrations of PBDEs in indoor dusts sampled in this study, and compares them with selected relevant studies elsewhere. Although the preliminary nature of this survey is acknowledged, comparison of our data with

TABLE 4. Summary of Estimates of Exposure (ng day-1) of UK Adults and Toddlers to ΣBDE via Dust Ingestion, Inhalation, and Diet, and Relative Significance (%) of Each Pathway adult

toddler (6-24 months) d-1)

air food dust (mean) dust (high) Σ(mean dust ingestion) Σ(high dust ingestion)

5th percentile

median

0.18 90.5 0.07 1.78 90.8 92.5

0.82 90.5 0.36 8.71 91.7 100.0

intake (ng average 95th percentile 2.1 90.5 0.9 21.5 93.5 114.1

5th percentile

median

average

95th percentile

0.03 51.6 0.98 3.56 52.6 55.2

0.16 51.6 4.8 17.4 56.6 69.2

0.4 51.6 11.8 43.1 63.8 95.1

1.7 51.6 32.3 117.3 85.6 170.6

0.3 91.2 8.5

0.6 80.8 18.6

2.0 60.3 37.7

0.2 74.6 25.2

0.4 54.3 45.3

1.0 30.2 68.8

8.8 90.5 2.4 58.7 101.7 158

% contribution mean dust ingestion scenario air food dust

0.2 99.7 0.1

0.9 98.7 0.4

2.3 96.8 1.0

8.7 88.9 2.4

0.1 98.0 1.9

high dust ingestion scenario air food dust

0.2 97.9 1.9

0.8 90.5 8.7

1.9 79.3 18.9

other studies shows concentrations in our dust samples to be slightly higher than those in a recent study of dusts from other European countries (33), but on average 20 times lower than those found in dusts from Ottawa (20). The median ratio of congener 47:99 in the dust samples is lower than that in air, at 0.51. This suggests that following volatilization from products treated with the commercial formulations (which favors the lower brominated congener 47), the higher brominated congener 99 undergoes preferential sorption to dust. Exposure to PBDEs via Inhalation and Dust Ingestion. Using the same exposure scenarios and assumptions as employed to derive estimates of inhalation exposure to ΣPCB, we have estimated inhalation exposure to ΣBDE (Table 4). The PUF disk air samplers employed likely underestimate concentrations of the higher brominated congeners as these are more associated with the particulate phase. Hence, our estimates of inhalation exposure to higher brominated congeners are likely underestimates also. Recently, it was suggested that ingestion of indoor dust constitutes the principal pathway of exposure to PBDEs in Canada (13). Furthermore, a strong correlation was reported between PBDE concentrations in human milk and dust from donor homes in the greater Boston area (34). Hence, we have made a preliminary evaluation of the likely magnitude of external exposure via dust ingestion to PBDEs to the inhabitants of the West Midlands. We have assumed 100% absorption of intake, average adult and toddler dust ingestion figures of 4.16 and 100 mg day-1, and high dust ingestion figures for adults and toddlers of 55 and 200 mg day-1 (20). We have then estimated various plausible dust ingestion exposure scenarios, using 5th percentile, median, average, and 95th percentile concentrations in our dust samples. As with our estimates of inhalation exposure, the range of exposure estimates via dust ingestion are only indicative of the likely range within the population. Table 4 summarizes these estimates of adult and toddler exposure to ΣBDE via dust ingestion. It also summarizes the significance of exposure of both adults and toddlers via inhalation and dust ingestion both relative to each other and to exposure via the diet. UK adult dietary exposure is assumed to be 90.5 ng ΣBDE day-1 (12), with that of toddlers assumed to be 57% of that of adults as described for PCBs. In summary, Table 4 shows that while inhalation is unlikely to make a major contribution to overall UK exposure to ΣBDE

5.6 57.3 37.1

0.1 93.5 6.5

(up to 9% for adults, and 2% for toddlers), dust ingestion makes an importantsbut not an overwhelmingscontribution (up to 37% for adults, and 69% for toddlers). This is in marked contrast to two similar exposure exercises, both of which suggested that dust ingestion is a major, and in some instances the overwhelmingly most important, pathway of exposure to PBDEs for Canadian adults and toddlers (13, 20). For example, comparison of our exposure estimates under the “average dust ingestion, median concentration” scenario for both adults and toddlers (0.36 and 4.79 ng ΣBDE day-1, respectively), with Canadian estimates under identical exposure scenarios (7.5 and 99 ng ΣBDE day-1 for adults and toddlers, respectively (20)) suggest UK exposure via dust ingestion to beswhile appreciablesan order of magnitude below that of North Americans. In contrast, the limited data currently available suggest UK dietary exposure to be approximately a factor of 2 higher than Canadian dietary exposure (12). While the preliminary nature of our current database on PBDEs in dust must be stressed and further, more detailed study is required, this study suggests that differences in dust contamination may be the principal cause of the intercontinental differences in human body burdens of PBDEs (15). Comparison of our UK estimates of overall UK exposure to ΣBDE with those for Canadians (20) suggests the range of exposures experienced by the UK population to be about a factor of fifteen narrower than those experienced by Canadians. While this appears to be inconsistent with the similar range of ΣBDE concentrations observed in studies of UK and U.S. human milk (35, 36), it is important to note that Table 3 does not consider the impact of variable dietary exposures, which our previous UK duplicate diet studies have shown to range over at least a factor of 6.4 (12), and which will exert a greater relative influence on exposures of the UK population than North Americans. When such variations in dietary exposure are taken into account, the likely range of overall ΣBDE exposures experienced by UK adults and toddlers widens considerably. This study significantly augments our knowledge of the extent of contamination of both indoor air and dusts found in a range of commonly frequented indoor microenvironments in urban areas. On the evidence presented here, indoor environments currently constitute an appreciable source of UK human exposure to both PCBs and PBDEs. Our data suggest that the order-of-magnitude lower body burdens of VOL. 40, NO. 15, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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PBDEs in Europeans compared to North Americans is likely due to the much higher exposures via dust ingestions experienced by the latter population. Urgent action is required to further extend our knowledge of the impact that contamination of indoor environments with PCBs, PBDEs, and related compounds may exert on both current and future human exposure.

(18) Gevao, B.; Al-Bahloul, M.; Al-Ghadban, A. N.; Ali, L.; Al-Omair, A.; Helaleh, M.; Al-Matrouk, K.; Zafar, J. Polybrominated Diphenyl Ethers in indoor air in Kuwait: Implications for human exposure. Atmos. Environ. 2006, 40, 1419-142.

Acknowledgments

(20) Wilford, B. H.; Shoeib, M.; Harner, T.; Zhu, J.; Jones, K. C. Polybrominated Diphenyl Ethers in Indoor Dust in Ottawa, Canada: Implications for Sources and Exposure. Environ. Sci. Technol. 2005, 39, 7027-7035.

We gratefully acknowledge the provision of studentships from the Iranian Ministry of Health and Medical Education (S.H.) and the National Council of Science and Technology Mexico (CONACYT) (C.I.).

Supporting Information Available Statistical summaries of concentrations of individual PBDE congeners and PCB homologue groups. This material is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) Harrad, S. J.; Sewart, A. P.; Alcock, R.; Boumphrey, R.; Burnett, V.; Duarte-Davidson, R.; Halsall, C.; Sanders, G.; Waterhouse, K.; Wild, S. R.; Jones, K. C. Polychlorinated biphenyls (PCBs) in the British environment: sinks, sources and temporal trends. Environ. Pollut. 1994, 85, 131-147. (2) Food Standards Agency. Dioxins and PCBs in the UK Diet: 1997 Total Diet Study; Food Survey Information Sheet Number 04/ 00; Food Standards Agency UK, 2000. (3) Food Standards Agency. Dioxins and Dioxin-Like PCBs in the UK Diet: 2001 Total Diet Study Samples; Food Survey Information Sheet 38/03; Food Standards Agency UK, 2003. (4) Harrad, S.; Wang, Y.; Sandaradura, S.; Leeds, A. Human dietary intake and excretion of dioxin-like compounds. J. Environ. Monit. 2003, 5, 224-228. (5) Currado, G. M.; Harrad, S. A Comparison of Polychlorinated Biphenyl Concentrations in Indoor and Outdoor Air and the Potential Significance of Inhalation as a Human Exposure Pathway. Environ. Sci. Technol. 1998, 32, 3043-3047. (6) Heinzow, B. G. J.; Mohr, S.; Ostendorp, G.; Kerst, M.; Ko¨rner, W. Dioxin-like PCB in indoor air contaminated with different sources. Organohalogen Compd. 2004, 66, 2470-2475. (7) Kohler, M.; Tremp, J.; Zennegg, M.; Seiler, C.; Minder-Kohler, S.; Beck, M.; Lienemann, P.; Wegmann, L.; Schmid, P. Joint Sealants: An Overlooked Diffuse Source of Polychlorinated Biphenyls in Buildings. Environ. Sci. Technol. 2005, 39, 19671973. (8) Alcock, R. E.; Sweetman, A. J.; Prevedouros, K.; Jones, K. C. Understanding levels and trends of BDE-47 in the UK and North America: an assessment of principal reservoirs and source inputs. Environ. Int. 2003, 29, 691-698. (9) McDonald, T. A. A perspective on the potential health risks of PBDEs. Chemosphere 2002, 46, 745-755. (10) Bromine Science Environmental Forum. http://www.bsef.com (accessed January 2004). (11) Renner, R. Increasing levels of flame retardants found in North American environment. Environ. Sci. Technol. 2000, 34, 452A453A. (12) Harrad, S.; Wijesekera, R.; Hunter, S.; Halliwell, C.; Baker, R. A Preliminary Assessment of UK Human Dietary and Inhalation Exposure to Polybrominated Diphenyl Ethers. Environ. Sci. Technol. 2004, 38, 2345-2350. (13) Jones-Otazo, H.; Clarke, J. P.; Diamond, M. L.; Archbold, J. A.; Ferguson, G.; Harner, T.; Richardson, G. M.; Ryan, J. J.; Wilford, B. Is House Dust the Missing Exposure Pathway for PBDEs? An Analysis of the Urban Fate and Human Exposure to PBDEs. Environ. Sci. Technol. 2005, 39, 5121-5130. (14) Stapleton, H. M.; Dodder, N. G.; Offenberg, J. H.; Schantz, M. M.; Wise, S. A. Polybrominated diphenyl ethers in house dust and clothes dryer lint. Environ. Sci. Technol. 2005, 39, 925-931. (15) Hites, R. A. Polybrominated Diphenyl Ethers in the Environment and in People: A Meta-Analysis of Concentrations. Environ. Sci. Technol. 2004, 38, 945-956. (16) Darnerud, P. O.; Eriksen, G. S.; Jo´hannesson, T.; Larsen, P. B.; Viluksela, M. Polybrominated Diphenyl Ethers: Occurrence, Dietary Exposure, and Toxicology. Environ. Health Perspect. 2001, 109 (Suppl 1), 49-68. (17) Harrad, S.; Diamond, M. Exposure to PBDEs and PCBs: Current and Future Scenarios. Atmos. Environ. 2006, 40, 1187-1188. 4638

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ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 40, NO. 15, 2006

(19) Wilford, B. H.; Harner, T.; Zhu, J.; Shoeib, M.; Jones, K. C. A Passive Sampling Survey of Polybrominated Diphenyl Ether Flame Retardants in Indoor and Outdoor Air in Ottawa, Canada. Environ. Sci. Technol. 2004, 38, 5312-5318.

(21) Covaci, A.; Voorspoels, S.; de Boer, J. Determination of brominated flame retardants, with emphasis on polybrominated diphenyl ethers (PBDEs) in environmental and human samples - a review. Environ. Int. 2003, 29, 735-756. (22) Currado, G. M.; Harrad, S. Factors Influencing Atmospheric Concentrations of Polychlorinated Biphenyls in Birmingham, U.K. Environ. Sci. Technol. 2000, 34, 78-82. (23) Harrad, S.; Mao, H. Atmospheric PCBs and Organochlorine Pesticides in Birmingham, UK: Concentrations, Sources, Temporal and Seasonal Trends. Atmos. Environ. 2004, 38, 14371445. (24) Kim, Y. M.; Harrad, S.; Harrison, R. M. Concentrations and Sources of VOCs in Urban Domestic and Public Microenvironments. Environ. Sci. Technol. 2001, 35, 997-1004. (25) Hazrati, S.; Harrad, S. Implications of Passive Sampling Derived Concentrations of Airborne PCBs and PBDEs in Urban Indoor Microenvironments. Organohalogen Compd. 2005, 67, 10331036. (26) Ayris, S.; Currado, G. M.; Smith, D.; Harrad, S. GC/MS procedures for the Determination of PCBs in Environmental Matrices. Chemosphere 1997, 35, 905-917. (27) Kemmelein, S.; Hahn, O.; Jann, O. Emissions of Organophosphate and Brominated Flame Retardants from Selected Consumer Products and Building Materials. Atmos. Environ. 2003, 37, 5485-5493. (28) Hoh, E.; Hites, R. A. Brominated Flame Retardants in the Atmosphere of the East-Central United States. Environ. Sci. Technol. 2005, 39, 7794-7802. (29) Sjo¨din, A.; Jakobsson, E.; Kierkegaard, A.; Marsh, G.; Sellstro¨m, U. Gas chromatographic identification and quantification of polybrominated diphenyl ethers in a commercial product, Bromkal 70-5DE. J. Chromatogr. A 1998, 822, 83-89. (30) Gearhart, J.; Posselt, H. Toxic at any speed. Chemicals in cars and the need for safe alternatives. The Ecology Center: Ann Arbor, MI, January 2006. (31) Vardoulakis, S. Personal communication. (32) Wearne, S.; Harrison, N.; Gem, M.; Startin, J.; Wright, C.; Kelly, M.; Robinson, C.; White, S.; Hardy, D.; Edinburgh, V. Time trend in human dietary exposure to PCDDs, PCDFs and PCBs in the UK. Organohalogen Compd. 1996, 30, 1-6. (33) Fabrellas, B.; Martı´nez, M. A.; Ramos, B.; Ruiz, M. L.; Navarro, I.; de la Torre, A. Results of European Survey based on PBDEs Analysis in Household Dust. Poster presented at 25th International Symposium on Halogenated Persistent Organic Pollutants (Dioxin 2005), Toronto, Canada, August 21-26, 2005. (34) Wu, N.; Webster, T.; Herrmann, T.; Paepke, O.; Tickner, J.; Hale, R.; Harvey, E.; La Guardia, M.; Jacobs, E. Associations of PBDE Levels in Breast Milk with Diet and Indoor Dust Concentrations. Organohalogen Compd. 2005, 67, 654-657. (35) Kalantzi, O. I.; Martin, F. L.; Thomas, G. O.; Alcock, R. E.; Tang, H. R.; Drury, S. C.; Carmichael, P. L.; Nicholson, J. K.; Jones, K. C. Different Levels of Polybrominated Diphenyl Ethers (PBDEs) and Chlorinated Compounds in Breast Milk from Two U.K. Regions. Environ. Health Perspect. 2004, 112, 1085-1091. (36) Schecter, A.; Pavuk, M.; Pa¨pke, O.; Ryan, J. J.; Birnbaum, L.; Rosen, R. Polybrominated Diphenyl Ethers (PBDEs) in U. S. Mothers’ Milk. Environ. Health Perspect. 2003, 111, 1723-1729.

Received for review April 15, 2006. Revised manuscript received June 8, 2006. Accepted June 13, 2006. ES0609147