Environ. Sci. Technol. 2010, 44, 4198–4202
Dust from U.K. Primary School Classrooms and Daycare Centers: The Significance of Dust As a Pathway of Exposure of Young U.K. Children to Brominated Flame Retardants and Polychlorinated Biphenyls S T U A R T H A R R A D , * ,† E M M A G O O S E Y , † JENNIFER DESBOROUGH,† M O H A M E D A B O U - E L W A F A A B D A L L A H , †,‡ LAURENCE ROOSENS,§ AND A D R I A N C O V A C I §,| 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, Department of Analytical Chemistry, Faculty of Pharmacy, Assiut University, 71526 Assiut, Egypt, Toxicological Centre, Department of Pharmaceutical Sciences, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk, Belgium, and Laboratory for Ecophysiology, Biochemistry and Toxicology, Department of Biology, University of Antwerp, Groenenborgerlaan 171, 2020 Antwerp, Belgium
Received March 8, 2010. Revised manuscript received April 22, 2010. Accepted April 23, 2010.
Polybrominated diphenyl ethers (PBDEs), hexabromocyclododecanes (HBCDs), tetrabromobisphenol-A (TBBP-A), and polychlorinated biphenyls (PCBs) were measured in floor dust from U.K. child daycare center and primary school classrooms (n ) 43, 36 for PCBs). Concentrations of HBCDs exceeded significantly (p < 0.05) those reported previously for U.K. houses and offices, while those of TBBP-A exceeded significantly thoseinU.K.carsandoffices.PCBconcentrationswerestatistically indistinguishable from those in U.K. house dust but lower than in U.S. classroom dust, while BDEs 47, 99, 100, 153, 196, 197, 203, and 209 in classrooms were significantly below concentrations in U.K. cars. Exposure of young U.K. children via classroom dust exceeds that of U.K. adults via office dust for all contaminants monitored. Overall dust exposure of young U.K. children was estimated including car, classroom, and house dust. Exposure to TBBP-A was well below a U.K. healthbased limit value (HBLV). Though no HBLVs exist for non-dioxinlike PCBs and HBCDs; dust exposure to PCBs fell well below U.K. dietary and inhalation exposure. Contrastingly, a high-end estimate of HBCD dust exposure exceeded U.K. dietary exposure substantially. Moreover, high-end estimates of dust
* Corresponding author e-mail:
[email protected]; phone: +44 121 414 7298; fax: +44 121 414 3078. † University of Birmingham. ‡ Assiut University. § Department of Pharmaceutical Sciences, University of Antwerp. | Department of Biology, University of Antwerp. 4198
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exposure to BDE-99 and BDE-209 (4.3 and 13000 ng/kg bw/ day, respectively) exceeded HBLVs of 0.23-0.30 and 7000 ng/ kg bw/day respectively.
Introduction Brominated flame retardants (BFRs) have found substantial use in consumer and commercial applications. In 2001, the last year for which figures are available, global production of tetrabromobisphenol-A (TBBP-A), hexabromocyclododecane (HBCD), decabromodiphenyl ether (Deca-BDE), octabromodiphenyl ether (Octa-BDE), and pentabromodiphenyl ether (Penta-BDE) amounted to 119,700, 16,700, 56,100, 3,790, and 7,500 t, respectively (1). While growing understanding of the potential human health effects of some BFRs has contributed to various jurisdictions acting to ban or limit severely manufacture and use of PBDEs, very few health based exposure standards exist. This situation is exemplified by the current position of the U.K. Committee on Toxicity (COT) that there is not yet sufficient evidence to permit setting a tolerable daily intake (TDI) for polybrominated diphenyl ethers (PBDEs) (2). Recently however, a preliminary Health Based Limit Value (HBLV) for BDE-99 was derived by Netherlands researchers (3). This HBLV (0.23-0.30 ng/kg bw/day) is driven by impaired spermatogenesis which was considered the most sensitive endpoint (3). Also pertinent given its predominance in dust from U.K. cars, homes, and offices (4) is the reference dose (RfD) of daily oral exposure to BDE-209 of 7 µg/kg bw/day (5). This RfD is that considered by the United States Environmental Protection Agency’s Integrated Risk Information System (IRIS) Toxicological Evaluations to be without appreciable risk of deleterious effects. With respect to HBCD, concerns have been expressed about its adverse health impacts in laboratory animals (6-9). Hence, while production continues and no recognized healthbased standard for HBCD exists, it is under active consideration for listing under the Stockholm Convention, and the European Chemicals Agency (ECHA) has identified it as a substance of very high concern under the REACH regulation because of its PBT properties. ECHA prioritized R-, β-, and γ-HBCD for inclusion in Annex XIV, the list of substances subject to authorization (10). Furthermore, TBBP-A has been identified as an endocrine disrupter because of its structural similarity to 17-estradiol and thyroxine (T4). It also displays a high potency to bind to human transthyretin (11, 12) and immunotoxicity (13). The U.K. COT has published a TDI for TBBP-A of 1 mg/kg dw/day (14). Such health concerns are compounded by exposure assessments that suggest toddlers receive considerable exposure to BFRs, like PBDEs and HBCDs, via ingestion of indoor dust (4, 15). However, while assessments of exposure via dust ingestion for adults include dust from homes, cars, and offices; those conducted to date for toddlers and young children have included only homes and cars (4, 15). Indeed to our knowledge there are no data whatsoever on concentrations of BFRs in dust from classrooms in child daycare centers and primary schools. While production and use of polychlorinated biphenyls (PCBs) ceased in industrialized countries, including the U.K., in the late 1970s, their widespread deployment and use in applications like permanently elastic sealants in buildings (16), means that their concentrations in indoor air may exceed considerably those in outdoor air (17). While health concerns have hitherto been associated primarily with exposure to dioxinlike congeners like PCB 126, evidence grows of adverse neurodevelopmental impacts of non-dioxin-like PCBs (18-20). 10.1021/es100750s
2010 American Chemical Society
Published on Web 05/04/2010
Moreover, while exposure via ingestion of house dust has been shown recently to constitute a relatively minor pathway of exposure to U.K. children (21), reports of elevated concentrations of PCBs in window caulk from schools in New England (22) and New York City (23), as well as concentrations of PCBs in dust from U.S. child daycare centers (24, 25), raise concerns about similar exposures in U.K. classrooms. This paper reports the concentrations of a range of trithrough deca-BDEs, HBCDs, TBBP-A, and PCBs in indoor dust sampled from 43 classrooms frequented by young children (age range ∼1-6 years). This is the first report worldwide of concentrations of BFRs in dust from school classrooms and child daycare centers. Coupled with data from previous studies of concentrations of these chemicals in dust from U.K. homes and cars, these data are used to assess the potential exposure of British children to such contaminants via dust ingestion.
Methods Sampling. Dust samples (n ) 43) were collected from classrooms in child daycare centers and primary schools in the West Midlands of the U.K., during winter 2007-spring 2008. Samples were collected using a portable vacuum cleaner, to which a sock with a 25 µm mesh size (Allied Filter Fabrics Ltd., Australia) was inserted into the nozzle of the device to retain dust. In contrast to our previous studies of homes and offices, where a standard specified proportion of the floor surface area of the room was sampled (4, 21, 26), and in recognition of the tendency of toddlers and young children to explore all parts of a room, the entire floor surface of each classroom was vacuumed thoroughly (typically for 4 min, depending on classroom size). Sampled room floor surface areas and dust loadings (mass of dust per unit floor surface area) ranged between 6-12 m2 and 0.025-0.3 g/m2. The approach taken in this study is more akin to sampling protocols employed by other research groups (27, 28). However, as noted previously, the extent to which such differences in sampling methods influence BFR concentrations detected in dust samples is not known (29). The socks containing the samples were placed in resealable polyethylene bags for transportation. The samples were sieved through a 500 µm mesh and stored in the dark at 4 °C until extraction. Analysis. Concentrations of PCBs (congeners 28/31, 52, 101, 118, 138, 153, and 180), R-, β-, γ-HBCD, and TBBP-A were determined at the University of Birmingham via GCEI/MS (PCBs) and LC-MS/MS (HBCDs and TBBP-A). Because of limited sample masses collected in some classrooms and the separate extraction for PCB analysis, PCBs were measured in 36 samples only. Concentrations of PBDEs 28, 47, 66, 85, 99, 100, 153, 154, 183, 196, 197, 203, and 209 were measured at the University of Antwerp via GC-ECNI/MS. Details of the methods used for extraction, clean up, and analysis of these compounds are described elsewhere (4, 21, 26), but are also provided in the Supporting Information. Statistical Analysis. Statistical analysis of the data was conducted using Microsoft Excel for Mac 2008 to generate descriptive statistics, with all other statistical procedures conducted using PASW Statistics for Mac. As a first step, the distribution of concentrations of each congener within the data set for each microenvironment category was evaluated for normality using both ShapirosWilk and Kolmogorovs Smirnov tests. The results, combined with visual inspection of frequency diagrams, revealed that data in each microenvironment category (cars, classrooms, homes, and offices) was lognormally distributed. Hence, concentrations were log-transformed before performing ANOVA or t test comparison of concentrations of individual congeners detected in dusts from each category of microenvironment studied. In all instances, where concentrations were below the detection limit, the
concentration was assumed to equal half of the detection limit for the purposes of statistical analyses.
Results and Discussion Table 1 summarizes the concentrations of target compounds in the analyzed samples. Concentrations of PCBs fall within the range reported for U.K. house dust (21), although two samples displayed extremely high concentrations (490 and 120 ng g-1) of PCB 28/31 that far exceed the maximum of 39 ng g-1 observed in samples of Canadian, New Zealand, U.K., and U.S. house dust collected in 2006. However, Table 1 also shows concentrations of PCBs in dust collected in 1997 from U.S. child daycare centers (n ) 10) to exceed substantially those in our U.K. classroom samples. Such higher concentrations in the U.S. are consistent with our previous findings of higher PCB concentrations in U.S. compared to U.K. house dust (21). In general, the predominant congeners in U.K. classroom dust are PCBs 28/31 and 52. This is consistent with the pattern observed in U.K. house dust (21) that differs markedly from that observed in Canadian and U.S. house dust and U.S. classroom dust where mid to high-molecular weight congeners predominate (21, 24, 25). While we have no definitive information on the Aroclor formulations used in the classrooms monitored, this suggests Aroclor 1242 to be the predominant source to the U.K. classroom samples, with Aroclor 1254-1260 playing a more important role in North America (21). Similar to previous studies of U.K. car, house, and office dust (4), the PBDE congener pattern is dominated by BDE209. Coupled with comparatively low concentrations of Σtrihexa-BDEs and BDE-183, this indicates far more substantial use of Deca-BDE than either Penta-BDE or Octa-BDE in the classrooms studied. Interestingly, while concentrations of BDE-209 in classroom dust are, like those in dust from other U.K. microenvironments (4), highly positively skewed (range 49-88,000 ng g-1), the range was not as great as those detected previously in samples of dust from U.K. cars and homes, where concentrations of BDE-209 reached up to 2.6 mg g-1 (4). Concentrations of ΣHBCDs in classroom dust are positively skewed (range 72-89,000 ng g-1), with levels in nine samples exceeding 10,000 ng g-1. The average ( σn-1 relative abundance of HBCD diastereomers is R- (37 ( 15%), β- (14 ( 4.9%), and γ-HBCD (48 ( 14%). A similar shift from the predominance of γ-HBCD in commercial formulations toward a greater proportion of R-HBCD has been observed previously in dust from other indoor environments from Canada, the U.K., and the U.S. (15, 26). This shift in diastereomer pattern is likely to result partly from the temperatures (160 °C-220 °C depending on the application) required to incorporate the commercial HBCD formulation into treated products such as expanded or extruded polystyrene. This has been shown to induce a marked shift from the γ- to the R-diastereomer (30). Such a shift is reflected presumably in the treated products and therefore in indoor dust. In addition, a shift from γ- to R-HBCD has been shown to occur via photoisomerization (31). Concentrations of TBBP-A in classroom dust span a wide range (17-1,400 ng g-1), with concentrations in six samples exceeding the highest concentration reported in dust from U.K. cars, homes, offices, and public microenvironments (15). Comparison with Concentrations in Other U.K. Microenvironments. Where possible, we compared concentrations detected in this study with those reported previously in dust samples from U.K. homes and cars, both microenvironments of potential relevance to toddlers and young children, as well as offices (4, 15, 21). t test comparison of log-transformed concentrations of both individual PCBs and ΣPCBs in this study with those reported previously for U.K. VOL. 44, NO. 11, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 1. Concentrations (ng g-1) in Indoor Dust Samples from U.K. Primary School and Daycare Center Classrooms of BFRs (n = 43) and PCBs (n = 36) and Comparison with Previous Relevant Studies
compound
minimum
5th percentile
BDE 28 BDE 47 BDE 66 BDE 100 BDE 99 BDE 85 BDE 154 BDE 153 Σtrihexa-BDEs BDE 183 BDE 197 BDE 203 BDE 196 BDE 209 ΣPBDEs R-HBCD β-HBCD γ-HBCD ΣHBCDsb TBBP-A PCB 28/31 PCB 52 PCB 101 PCB 118 PCB 138 PCB 153 PCB 180 ΣPCBsc
