- Email: [email protected]
Polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in human liver and adipose tissue samples from Belgium
Chemosphere 73 (2008) 170–175
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in human liver and adipose tissue samples from Belgium Adrian Covaci a,b,*, Stefan Voorspoels a, Laurence Roosens a, Werner Jacobs c, Ronny Blust b, Hugo Neels a a
Toxicological Centre, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk-Antwerp, Belgium Ecophysiology, Biochemistry and Toxicology Group, Department of Biology, University of Antwerp, Groenenborgerlaan 171, 2020 Antwerp, Belgium c Department of Forensic Medicine, University Hospital of Antwerp, Wilrijkstraat 10, 2650 Antwerp, Belgium b
a r t i c l e
i n f o
Article history: Received 2 December 2007 Received in revised form 23 February 2008 Accepted 26 February 2008 Available online 11 April 2008 Keywords: Polybrominated diphenyl ethers Polychlorinated biphenyls Human liver Human adipose tissue Age and gender dependency Belgium
a b s t r a c t Levels of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) were measured in paired human adipose tissue and liver samples (n = 25) from Belgium. Average concentrations and standard deviation of sum PBDEs (congeners 28, 47, 99, 100, 153, 154 and 183) were 5.3 ± 3.0 (range 1.4–13.2) and 3.6 ± 2.1 (range 1.0–10.0) ng g1 lipid weight (lw) in adipose tissue and liver, respectively. These concentrations were similar to reported PBDE data from Belgium and were at the lower end of the concentration range reported elsewhere in the world. In both tissues under study, BDE 153 and BDE 47 were the most abundant PBDE congeners, contributing approximately 35% and 25% to the total PBDE content. Average concentrations and range of PCBs (sum of 23 congeners) were 490 (range 70–1130) and 380 (range 90–1140) ng g1 lw in adipose tissue and liver, respectively. No correlation between age and concentrations of PBDEs could be found (r = 0.04), while PCB concentrations correlated significantly with age (r = 0.62, p < 0.01, for the sum PCBs; r = 0.64, p < 0.01 for PCB 153 alone). Factors, such as exposure pathways (food, dust and air), rates of bioaccumulation, metabolism and elimination, influence the concentrations of PBDEs differently than those of PCBs in humans. Ó 2008 Elsevier Ltd. All rights reserved.
1. Introduction The massive use of polybrominated diphenyl ethers (PBDEs) as flame retardants in thermoplastics (e.g. computer and TV housing), textiles, foams, interiors of cars, buses and airplanes is related to strict fire regulations. As a consequence, these products are extensively present in our daily life. Concurrent with the increasing use, environmental levels of PBDEs have risen since their first application. Spillage and emission during production and use, release from the consumer products when used and also disposal at the end-oflife of the consumer products, account for this phenomenon. These compounds are chemically and biologically persistent, lipophilic, bioaccumulate in fatty tissues and biomagnify throughout food chains (Law et al., 2006). Recently, the constituents of Penta-BDE technical mixtures (mainly consisting of tri- to hexa-BDE congeners) have been proposed as persistent organic pollutants (POPs) candidates since they
* Corresponding author. Address: Toxicological Centre, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk, Belgium. Tel.: +32 3 820 2704; fax: +32 3 820 2722. E-mail address: [email protected] (A. Covaci). 0045-6535/$ - see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2008.02.059
embody all characteristics of the Stockholm convention definition of POPs: bioaccumulation, toxicity, persistency and long-range transport potential (WWF, 2005). Penta-BDE mixtures, as well as the octa-BDE mixture, have been banned in 2004 in the European Union (EEC, 2003). Although the human exposure pathways to these chemicals through food and dust is increasingly being investigated (Schecter et al., 2005a; Stapleton et al., 2005; Harrad and Diamond, 2006a; Voorspoels et al., 2007; Harrad et al., 2008), any causality with the observed levels in human tissues is still far from understood. Recently, Wu et al. (2007) have reported a significant positive correlation between PBDEs in human milk and in dust from the homes of donors (n = 11), suggesting that dust might be an important exposure pathway for humans. The existing database with levels of PBDEs in humans is yet limited and contains mostly data from serum and milk (Thomsen et al., 2002; Sjödin et al., 2003, 2004; Gill et al., 2004; Schecter et al., 2005b). The present study aims to expand the knowledge concerning human exposure to PBDEs and PCBs in a Western European country and to discuss the extent of contamination and distribution of these contaminants two lesser studied tissues, namely liver and adipose tissue. Additionally, we explored the relationships between age or gender and levels of PBDEs and PCBs.
171
A. Covaci et al. / Chemosphere 73 (2008) 170–175
2. Materials and methods
2.3. Quality assurance
2.1. Sample collection
Retention times, ion chromatograms and relative abundance of the monitored ions were used as identification criteria. A deviation of ion abundance ratios within 15% of the mean values for calibration standards was considered acceptable. Quantification was based on five-point calibration curves in which either the sum of m/z = 79 and 81 for PBDEs or the most intense ion from the molecular cluster for PCBs were monitored. The method performance was assessed through rigorous internal quality control, which included daily check of calibration curves, regular analysis of method blanks and certified reference materials (BCR 350 – PCBs in mackerel oil and SRM 1945 – PCBs and PBDEs in whale blubber). For quality assurance and control, the laboratory participated in the Interlaboratory Comparison Exercise Program for Organic Contaminants in Marine Mammal Tissues organized by the National Institute of Standards and Technology (NIST, Gaithersburg, MD, USA). Results for individual PCB and PBDE congeners had a variation coefficient less than 15% from the target values (Kucklick et al., 2005, 2007). Since BDE 153 was found in unusual high proportion in several samples, special attention was given to its quality assurance. As previously reported by us (Korytar et al., 2005), BDE 153 was not co-eluted with other major PBDE congeners or with hexabromocyclododecane. Spiking experiments (n = 3) at 5 ng/g lw done in pork liver indicated a recovery of 86 ± 4%. Further, results for BDE 153 (together with BDE 47 and CB 153) obtained during participation in different interlaboratory exercises are given in Table 1. Some degradation probably arising from both the column and injector was observed for PBDE congeners with 8–10 bromine atoms and therefore, no further efforts were made to quantify these congeners, which might have been present in the samples. For PBDE and PCB congeners that were consistently measured in the procedural blanks, the mean blank value was subtracted. BDE 47 and BDE 99 were the only PBDE congeners detected in the blanks with mean value (SD) of 0.07 ± 0.03 ng and 0.06 ± 0.02 ng, respectively. The limit of quantification (LOQ) was calculated as 2 SD of the blank values and related to the sample intake, namely 0.2 and 5 g for adipose tissue and liver, respectively (Voorspoels et al., 2003). For PBDE (including BDE 153) and PCB congeners not detected in the blanks, the LOQ was calculated based on a instrumental signal-to-noise of 10. LOQs ranged between 0.03 and 0.2 ng g1 lipid weight (lw) for tri- to hepta-BDEs and between 1 and 4 ng g1 lw for tri- to octa-CBs.
The present study was approved by the Ethics Committees of the University Hospital of Antwerp and of the University of Antwerp. Paired human liver and adipose tissue (mesenteric fat) samples (n = 25) were collected between 2003 and 2005 during autopsy at the Department of Forensic Medicine of the University Hospital of Antwerp. Samples originated from 18 males and 7 females, who died from causes unrelated to environmental contaminants as far as could be determined. The mean age of the individuals was 37 years (range 9–70 years), while the mean weight was 75 kg (range 30–100 kg). No information about possible occupational exposure or dietary habits was available. Samples were collected in hexane-washed polyethylene recipients, frozen immediately and stored at 20 °C until analysis. 2.2. Chemical analysis The sample preparation and analysis method was previously validated and reported elsewhere in detail (Covaci et al., 2002; Voorspoels et al., 2003). Based on their reported abundance, the following contaminants (IUPAC numbering), were targeted for analysis: PBDE congeners (Nos. 28, 47, 99, 100, 153, 154 (co-eluting with BB 153), and 183) and PCB congeners (Nos. 28/31, 52, 74, 95, 99, 101, 105, 110, 118, 128, 138/163, 149, 153, 156, 170, 180, 183, 187, 194, 196, and 196). The individual PCB standards were obtained from Dr. Ehrenstorfer Laboratories (Augsburg, Germany) and PBDE standard mixtures were obtained from Wellington Laboratories (Guelph, Ontario, Canada). Hexane, acetone, dichloromethane, iso-octane (all pesticide grade) and concentrated sulfuric acid (analytical grade) were obtained from Merck (Darmstadt, Germany). Silica gel 60 (63–230 mesh) and anhydrous Na2SO4 (Merck) were heated at 150 °C for 24 h. Liver (3 g) or adipose tissue samples (0.3 g) were mixed with anhydrous Na2SO4, spiked with internal standards (BDE 77, BDE 128, CB 46 and CB 143) and Soxhlet-extracted with hexane:acetone (3:1, v/v). After gravimetrical lipid determination (on an aliquot), the extract was cleaned on acidified silica gel and the analytes were eluted with 15 ml n-hexane and 10 ml dichloromethane. The eluate was concentrated to near dryness and reconstituted in 100 ll iso-octane. The analyses were performed on an Agilent 6890 GC coupled with an Agilent 5973 MS. For PCBs, electron ionization (EI) mode was used in combination with a 30 m 0.25 mm 0.25 lm DB-1 capillary column (J&W Scientific, Folsom, CA, USA). For PBDEs, electron capture negative ionization (ECNI) mode was used in combination with a 12 m 0.18 mm 0.20 lm AT-5 column (Alltech, Lokeren, Belgium). Instrumental parameters were presented in detail by Covaci et al. (2002) and Voorspoels et al. (2003).
2.4. Data treatment Concentrations below the LOQ were assigned a value of f*LOQ, with ‘f ’ the proportion of measurements with levels above the LOQ (Voorspoels et al., 2002). This approach allowed data below LOQ to be used in the statistical data treatment. All statistical
Table 1 Values obtained by the Toxicological Center for BDE 47, BDE 153 and CB 153 in recent interlaboratory exercises BDE 47
AMAP 2007 Round 1 (ng/ml)
Round 3 (ng/ml)
NIST 2005 Unknown blubber (ng/g) SRM 1945 (ng/g)
BDE 153
Mean ± SD
Assigned
0.36 0.90 0.99 0.74 0.44 0.99
0.44 0.97 0.99 0.73 0.44 0.98
123 ± 4 36.4 ± 0.5
132 ± 17 39.6 ± 0.2
Mean ± SD 0.19 0.37 0.24 0.33 0.21 0.44 12.6 ± 0.4 7.5 ± 0.5
CB 153 Assigned 0.22 0.42 0.31 0.35 0.25 0.46 14.2 ± 1.6 8.3 ± 0.6
Mean ± SD
Assigned
1.93 3.47 1.17 3.12 3.91 1.09
1.9 3.1 1.2 3.0 3.9 1.1
1700 ± 40 211 ± 3
1830 ± 240 228 ± 10
172
A. Covaci et al. / Chemosphere 73 (2008) 170–175
analyses were performed using Statistica for Windows and GraphPad Instat version 3.06 for Windows. The normal distribution of concentrations of each congener within the dataset for tissue was evaluated using both Shapiro-Wilks and K–S tests. Since concentrations were normally distributed, a paired t-test was used to test differences between PBDE concentrations in liver and adipose tissue. No outliers were detected using Grubbs’ test. Pearson correlation coefficients were used to measure the strength of the association between age and the organohalogen concentrations. 3. Results and discussion 3.1. Residue levels of PBDEs and comparison with other data The average lipid percentage and standard deviation in liver was 7.0 ± 3.8% (range 3.3–16.9%), while the average lipid percentage in adipose tissue was 86.2 ± 6.1% (range 70.1–94.1%). All investigated PBDE congeners had a detection frequency higher than 80%, with BDE 153 and BDE 154/BB 153 being detected in all samples. The co-elution between BDE 154 and BB 153 could not be resolved on the AT-5 column used with the GC-ECNI/MS detection (Covaci et al., 2003, 2007). However, since BDE 154 is reported as a minor PBDE congener in humans (Covaci et al., 2002; Sjödin et al., 2003), we can assume that BB 153 had in these samples a higher contribution to the peak. This assumption is also strengthened by the fact that BB 153 was previously measured in serum of Belgian donors, while BDE 154 could not be detected in the same samples (Covaci and Voorspoels, 2005). Average concentration of sum PBDEs (sum of congeners 28, 47, 99, 100, 153) in Belgian liver samples was 3.6 ± 2.1 ng g1 lw, range 1.0–10.0 ng g1 lw (Table 2). For adipose tissue, average of sum PBDEs was 5.3 ± 3.0 ng g1 lw, range 1.4–13.2 ng g1 lw (Table 2). There were no outliers between the concentrations of PBDEs in the present study, while other studies have reported the presence of 5% values with high concentrations (Harrad and Porter, 2007). PBDE concentrations in the present study were similar to other reported concentrations in Belgium. Table 3 summarizes all available Belgian data in different human matrices (adipose tissue, milk and serum). Concentrations of PBDEs in Belgian adipose tissue and liver were similar to reported concentrations in tissues from other European countries (Strandman et al., 1999; Smeds and Saukko, 2003; Fernandez et al., 2007; Meneses et al., 1999; Meironyté Guvenius et al., 2001) and were notably lower than those in USA (Petreas et al., 2003; Johnson-Restrepo et al., 2005). In a Swedish survey, BDE 47 was measured in more than 400 individuals (Hardell
Table 2 Average concentrations (ng g1 lw), standard deviations, detection frequency and range (in brackets) of PBDE congeners in paired liver and adipose tissue samples from 25 Belgian individuals Detection frequency (%)
Adipose tissue (ng g1 lw)
Liver (ng g1 lw)
84 96 86 94 100 100 96
25 86 ± 6.1% (70–93%) 0.08 ± 0.06a 1.2 ± 1.1a 0.55 ± 0.47a 0.34 ± 0.33b 2.0 ± 1.8b 0.91 ± 0.59b 0.31 ± 0.12b
25 7.0 ± 3.8% (3.1–16.9%) 0.06 ± 0.04a 0.95 ± 0.83a 0.38 ± 0.36a 0.17 ± 0.20b 1.2 ± 1.4b 0.66 ± 0.43b 0.21 ± 0.12b
96 100 100 100 100 100 100 100
5.3 ± 3.0b (1.4–13.2) 9.7 ± 6.4a 21.0 ± 18.1a 83.0 ± 58.5a 131 ± 92b 40.5 ± 29.4a 98.1 ± 74.4a 10.5 ± 7.4b 26.1 ± 21.5b
3.6 ± 2.1b (1.0–10.0) 8.1 ± 8.0a 19.1 ± 18.1a 68.2 ± 55.0a 99.6 ± 80.4b 33.6 ± 24.3a 70.7 ± 57.8a 8.4 ± 6.8b 22.4 ± 19.0b
Sum 7 marker PCBs
334 ± 234b (47–798)
259 ± 205b (57–756)
Sum PCBs
490 ± 341b (68–1128)
377 ± 299b (89–1138)
N Lipids BDE BDE BDE BDE BDE BDE BDE
28 47 99 100 153 154/BB 153 183
Sum PBDEs CB CB CB CB CB CB CB CB
a b
99 118 138 + 163 153 170 180 183 187
Not significantly different. Significantly different (p < 0.05).
et al., 1998). Concentrations ranged between 0.5 and 4 ng g1 lw for most specimens with few persons having levels up to 100 ng g1 lw, indicating a large variation in the individual values. Adipose tissue samples from USA (Petreas et al., 2003) contained almost one order of magnitude higher concentrations of BDE 47 than our group, while even higher concentrations were measured in a newer study (Johnson-Restrepo et al., 2005). Concentrations of PBDEs in Japanese individuals were similar to or even lower than the Belgian concentrations (Choi et al., 2007; Kunisue et al., 2007). The differences between countries can be partially explained by the PBDE concentrations in the diet and especially in food items with higher contribution to the total PBDE intake (fish and meat products) (Voorspoels et al., 2007). However, other exposure pathways (e.g. indoor dust) have recently been identified for humans and their contribution to the total PBDE intake may be higher than
Table 3 Summary of PBDE concentrations (ng g1 lw) in Belgian human samples Sampling year
Sample type
Number of samples
Age
Sum PBDEsa (ng g1 lw) mean ± SD
% BDE 47 and BDE 153 to sum PBDEsa
Reference
1999–2004
Serum
11
n.a.
4.0 ± 1.6
31 and 43
2000–2001 2001 2002–2003
Milk Adipose tissue Pooled cord serum Pooled cord serum Serum Pooled serum Pooled serum Adipose tissue Liver Adipose tissue
14 20 4
26–38 19–77 0
2.9 4.7 ± 2.3 2.6 ± 0.4
59 and 15 31 and 52 49 and 25
0
2.2 ± 0.5
19 and 40
Covaci and Voorspoels 2005 Pirard et al., 2003 Covaci et al., 2002 Covaci and Voorspoels 2005 Roosens et al., 2007
19–63 14–15 50–65
3.4 ± 3.4 4.0 ± 0.4 4.6 ± 0.8 7.6 ± 8.9 3.6 ± 2.1 5.3 ± 3.0
36 32 24 22 26 22
Van Wouwe et al., 2004 Roosens et al., 2007 Roosens et al., 2007 Naert et al., 2006 This study This study
2002–2003 2003 2003–2004 2004–2005 2006 2002 and 2006 2002 and 2006
7 20 8 8 53 25 25
9–70 9–70
n.a. – Not available. a For each study, sum PBDEs = BDE 28 + 47 + 99 + 100 + 153 + 154 + 183.
and and and and and and
35 36 30 49 33 37
173
A. Covaci et al. / Chemosphere 73 (2008) 170–175
expected, at least for some groups, such as toddlers (Harrad et al., 2008). 3.2. Congener profile of PBDEs BDE 153 and BDE 47 were the most abundant PBDE congeners in liver, attributing for 33% and 26%, respectively. Similar to the liver, these two congeners were also the most abundant in adipose tissue, contributing to 37% and 22% to the total PBDE content (Fig. 1). High concentrations of BDE 153 were also seen in some samples from Spain (Meneses et al., 1999). BDE 153, as the dominant congener, was also observed in Japan (Choi et al., 2007; Kunisue et al., 2007), Spain (Meneses et al., 1999; Fernandez et al., 2007), Sweden (Meironyté Guvenius et al., 2001) and Belgium (Covaci et al., 2002). However, in the US, where greater amounts of technical Penta-BDE have been used, BDE 47 was by far the most predominant congener (60–70% of the total PBDEs) in human adipose tissue (Johnson-Restrepo et al., 2005; She et al., 2002). The difference in the profiles of PBDEs might be related to dietary habits or to temporal trends in the exposure to various PBDE technical mixtures. Furthermore, exposure through dust can have an important role in the resulting PBDE profiles present in the exposed individuals. The more lipophilic PBDE congeners (BDEs 100, 153, and 183) and the sum PBDEs had slightly higher statistically significant concentrations in adipose tissue compared to liver samples (Table 2 and Fig. 1). This can be explained by the relatively low metabolic activity of adipose tissue which is primarily used for energy (and pollutants) storage, while liver is an active tissue where contaminants are (eventually) more readily metabolized. Furthermore, the profiles of PBDE congeners in adipose tissue or liver differed from that in human milk (Meironyté et al, 1999) and in blood (Thomsen et al., 2002), with BDE 47 being by far the predominant PBDE congener (70% of the total sum PBDEs). Obviously, the PBDE profile in human tissues depends on the biotransformation and accumulation kinetics of each congener. Furthermore, the distribution of PBDE congeners between different tissues could be directly related to their half-lives, octanol-water partition coefficient, and the properties of the tissue (Inoue et al. 2006). The high percentage of BDE 153 observed in the investigated adipose tissue and liver samples might also be a consequence of its long half-life (Sjödin et al., 2003), and hence it is likely that
BDE 153 is more persistent and lipid-dependently accumulated in human bodies. The concentrations of individual PBDE congeners and sum PBDEs in liver and adipose tissue were significantly correlated (BDE 153: r = 0.93, p < 0.01, CA = 1.21CL + 0.52; BDE 47: r = 0.80, p < 0.01, CA = 1.04CL + 0.20; sum PBDEs: r = 0.74, p < 0.01, CA = 1.07CL + 1.46), CA and CL are the concentrations in adipose tissue and liver, respectively. The lower correlation coefficients obtained for sum PBDEs may be explained by the higher number of no-detects and lower values for PBDE congeners other than BDE 47 and 153, which greatly contributed to the sum PBDEs. 3.3. Residues levels and congener profiles of PCBs Several PCB congeners (CB 31/28, CB 52 and CB 128) were not detected in any sample, while other PCBs (CB 74, CB 95, CB 105, CB 149) had a detection frequency of less than 50%. These congeners were therefore not included in the statistical analysis. As expected, the persistent PCB congeners (CB 99, CB 118, CB 138/163, CB 153, CB 170, CB 183, CB 187 and CB 183) were detected in all samples and their descriptive parameters are given in Table 2. Average concentration of sum PCBs (sum of 23 PCB congeners in Belgian liver samples was 377 ± 299 ng g1 lw, range 89– 1138 ng g1 lw (Table 2). For adipose tissue, average of sum PCBs was 490 ± 341 ng g1 lw, range 68–1128 ng g1 lw (Table 2). Concentrations of PCBs measured in the present study were similar to previous studies in Belgian adipose tissue (Covaci et al. 2002; Naert et al. 2006). Concentrations of the sum 7 marker PCBs are also given as additional information in Table 2. As previously reported (Covaci et al., 2002; Naert et al., 2006), CB 153 and CB 180 were the major PCB congeners in human adipose tissue, similar to profiles reported from current European reports. Only few congeners (CB 153, CB 183 and CB 187) together with sum PCBs presented higher statistically significant concentrations in adipose tissue than in liver. Similar to PBDEs, this is probably due to the storage role of adipose tissue compared to the more active function of the liver. The concentrations of selected individual PCB congeners and sum PCBs in liver and adipose tissue were significantly correlated (CB 153: r = 0.72, p < 0.01, CA = 0.83CL + 49; CB 180: r = 0.62, p < 0.05, CA = 0.80CL + 42; sum PCBs: r = 0.69, p < 0.01, CA = 0.78CL + 195).
45
40 liver adipose tissue
35
% of sum PBDEs
30
25
20
15
10
5
0 BDE 28
BDE 47
BDE 100
BDE 99
BDE 154/BB153
BDE 153 *
BDE 183
Fig. 1. Congener profiles of PBDEs in adipose tissue and liver. Error bars represent 2 standard error (SE).
174
A. Covaci et al. / Chemosphere 73 (2008) 170–175
3.4. Age- and gender-dependent accumulation of PCBs and PBDEs Although it was suggested that the accumulation of PBDEs will increase with age (Hooper and McDonald, 2000), no clear evidence could be found in the scientific literature. In the present study, PBDE concentrations in human tissues (data from Covaci et al 2002; Van Wouwe et al., 2004; Naert et al., 2006; Roosens et al., 2007) could not be correlated with age (r = 0.004 for sum PBDEs; r = 0.04 for BDE 153). It suggests that human exposure to PBDEs is of recent nature and that these contaminants have not yet attained a steady-state in human tissues. In contrast, PCBs correlated significantly with age (r = 0.62, p < 0.01, for the sum PCBs; r = 0.64, p < 0.01 for PCB 153) (Fig. 2). Interestingly, some of the highest concentrations of PBDEs were measured in young specimens. This suggests that other sources than long-term dietary exposure (e.g. indoor dust or air) are also important for PBDEs. However, several additional factors, such as the rate of metabolism of PBDE congeners, the type of technical mixture to which the persons have been exposed, mother–child transfer, or life-stage specific exposures (e.g. toddlers), should also be taken into account when assessing the accumulation of PBDEs in humans. Similar to our results, Hardell et al. (1998) did not find an age-dependency for levels of BDE 47 in adipose tissue from 77 individuals (age range 28–85 years) with different types of cancer. They also concluded that the determinant factors for the concentrations of PBDEs in humans are most probably different than those for PCBs and other POPs. Although the number of samples for each gender group in the present study was limited, a tendency for higher concentrations of sum PBDEs in males was observed in adipose tissue, but not in liver (t-test, p < 0.05). This result is consistent with observations of Kunisue et al. (2007), which have measured higher concentra-
a conc PBDEs (ng/g lw)
45 40
r = 0.004
35 30 25 20 15 10 5 0 0
20
40
60
80
100
age
b conc sum PCBs (ng/g lw)
2500
r = 0.62, p < 0.01 2000 1500 1000 500 0 0
20
40
60
80
100
age Fig. 2. Pearson correlation between (a) age and sum PBDEs and (b) between age and sum PCBs in Belgian human samples. PBDE data are from Van Wouwe et al. (2004), Covaci and Voorspoels (2005), Naert et al. 2006 and from the present study.
tions of PBDEs in Japanese adipose tissue from males. Therefore, breast-feeding might play an important role in the elimination of PBDEs in females. The higher concentrations in males may also be attributable to the fact that their lifestyles may involve spending more time in contaminated microenvironments, such as offices and cars, which have been shown to be more contaminated than homes (Harrad et al, 2006b). In another study, no age or gender dependency with PBDE concentrations has been recently seen in individuals from New Zealand (Harrad and Porter, 2007). Contrarily, no gender differences for PCBs were observed in liver or in adipose tissue, which is also consistent with the findings of Kunisue et al. (2007). This indicates that the recent human exposure to PCBs is relatively constant, while the concentrations of PCBs in human tissues have already reached a steady-state. Acknowledgements Katrien Jorissen is thanked for the sample preparation. Dr. Caroline Naert is acknowledged for providing individual data on PCBs, PBDEs and age from her study (Naert et al., 2006). Adrian Covaci acknowledges a postdoctoral fellowship of the Funds for Scientific Research Flanders (FWO). The authors thank to the organizers of the BFR 2007 workshop in Amsterdam for allowing this study to be presented as a poster. References Choi, J., Fujimaki, S., Kitamura, K., Hashimoto, S., Ito, H., Suzuki, N., Sakai, S., Morita, M., 2007. Polybrominated dibenzo-p-dioxins, dibenzofurans, and diphenyl ethers in Japanese human adipose tissue. Environ. Sci. Technol. 37, 817–821. Covaci, A., de Boer, J., Ryan, J.J., Voorspoels, S., Schepens, P., 2002. Distribution of organobrominated and organochlorinated contaminants in Belgian human adipose tissue. Environ. Res. 88, 210–218. Covaci, A., Voorspoels, S., de Boer, J., 2003. Determination of BFRs with emphasis on PBDEs in environmental and human samples – a review. Environ. Int. 29, 735– 756. Covaci, A., Voorspoels, S., 2005. Optimization of the determination of polybrominated diphenyl ethers in human serum using solid-phase extraction and gas chromatography-electron capture negative ionization mass spectrometry. J. Chromatogr. B 827, 216–223. Covaci, A., Voorspoels, S., Ramos, L., Neels, H., Blust, R., 2007. Recent developments in the analysis of brominated flame retardants and brominated natural compounds. J. Chromatogr. A 1153, 145–171. EEC, 2003. Directive 2003/11/EC of the European Parliament and of the Council of 6 February 2003 amending for the 24th time Council directive 76/769/EEC relating to restrictions on the marketing and use of certain dangerous substances and preparations pentabromodiphenyl ether and octabromodiphenyl ether). Official J. L 042, 15/02/2003. Fernandez, M.F., Araque, P., Kiviranta, H., Molina-Molina, J.M., Rantakokko, P., Laine, O., Vartiainen, T., Olea, N., 2007. PBDEs and PBBs in the adipose tissue of women from Spain. Chemosphere 66, 377–383. Gill, U., Chu, I., Ryan, J.J., Feeley, M., 2004. Polybrominated diphenyl ethers: human tissue levels and toxicology. Rev. Environ. Contam. Toxicol. 183, 55–97. Hardell, L., Lindström, G., van Bavel, B., Wingfors, H., Sundelin, E., Liljegren, G., 1998. Concentrations of the flame retardant 2,20 ,4,40 -tetrabrominated diphenyl ether (BDE 47) in human adipose tissue in Swedish persons and the risk for nonHodgkin’s lymphoma. Oncol. Res. 10, 429–432. Harrad, S., Diamond, M., 2006a. New directions: exposure to polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs): current and future scenarios. Atmos. Environ. 40, 1187–1188. Harrad, S., Hazrati, S., Ibarra, C., 2006b. Concentrations of polybrominated diphenyl ethers in indoor air and dust and polychlorinated biphenyls in indoor air in Birmingham, United Kingdom: implications for human exposure. Environ. Sci. Technol. 40, 4633–4638. Harrad, S., Porter, L., 2007. Concentrations of polybrominated diphenyl ethers in blood serum from New Zealand. Chemosphere 66, 2019–2023. Harrad, S., Ibarra, C., Diamond, M., Melymuk, L., Robson, M., Douwes, J., Roosens, L., Dirtu, A.C., Covaci, A., 2008. Polybrominated diphenyl ethers in domestic indoor dust from Canada, New Zealand, United Kingdom and United States. Environ. Int. 34, 232–238. Hooper, K., McDonald, T.A., 2000. The PBDEs: an emerging environmental challenge and another reason for breast-milk monitoring programs. Environ. Health Perspect. 108, 387–392. Inoue, K., Harada, K., Takenaka, K., Uehara, S., Kono, M., Shimizu, T., Takasuga, T., Senthilkumar, K., Yamashita, F., Koizumi, A., 2006. Levels and concentration ratios of polychlorinated biphenyls and polybrominated diphenyl ethers in serum and breast milk in Japanese mothers. Environ. Health Perspect. 114, 1179–1185.
A. Covaci et al. / Chemosphere 73 (2008) 170–175 Johnson-Restrepo, B., Kannan, K., Rapaport, D.P., Rodan, B.D., 2005. Polybrominated diphenyl ethers and polychlorinated biphenyls in human adipose tissue from New York. Environ. Sci. Technol. 39, 5177–5182. Korytar, P., Covaci, A., de Boer, J., Gelbin, A., Brinkman, U.A.Th., 2005. Retention-time database of 126 PBDE congeners and 2 Bromkal technical mixtures on 7 capillary GC columns. J. Chromatogr. A 1065, 239–249. Kucklick, J.R., Pugh, R.S., Becker, P.R., Schantz, M.M., Wise, S.A., Rowles, T.K., 2005. Description and results of the 2003 NIST/NOAA interlaboratory comparison exercise program for organic contaminants in marine mammal tissues. NISTIR 7269, 157 p. Kucklick, J.R., Pugh, R.S., Becker, P.R., Schantz, M.M., Wise, S.A., Rowles, T.K. Description and results of the 2005 NIST/NOAA interlaboratory comparison exercise program for organic contaminants in marine mammal tissues. NISTIR 7410, 132 p. Kunisue, T., Takayanagi, N., Isobe, T., Takahashi, S., Nose, M., Yamada, T., Komori, H., Arita, N., Ueda, N., Tanabe, S., 2007. Polybrominated diphenyl ethers and persistent organochlorines in Japanese human adipose tissues. Environ. Int. 33, 1048–1056. Law, R.J., Allchin, C.R., de Boer, J., Covaci, A., Herzke, D., Lepom, P., Morris, S., Tronczynski, J., de Wit, C.A., 2006. Levels and trends of brominated flame retardants in the European environment. Chemosphere 64, 187– 208. Meneses, M., Wingfors, H., Schuhmacher, M., Domingo, J.L., Lindström, G., van Bavel, B., 1999. Polybrominated diphenyl ethers detected in human adipose tissue from Spain. Chemosphere 39, 2271–2278. Meironyté, D., Norén, K., Bergman, Å., 1999. Analysis of polybrominated diphenyl ethers in Swedish human milk. A time-related trend study, 1972–1997. J. Toxicol. Environ. Health 58, 329–341. Meironyté Guvenius, D., Bergman, Å., Norén, K., 2001. Polybrominated diphenyl ethers in Swedish human liver and adipose tissue. Arch. Environ. Contam. Toxicol. 40, 564–570. Naert, C., Piette, M., Bruneel, N., Van Peteghem, C., 2006. Occurrence of polychlorinated biphenyls and polybrominated diphenyl ethers in Belgian human adipose tissue samples. Arch. Environ. Contam. Toxicol. 50, 290–296. Petreas, M., She, J., Brown, F.R., Winkler, J., Windham, G., Rogers, E., Zhao, G.M., Bhatia, R., Charles, M.J., 2003. High body burdens of 2,20 ,4,40 tetrabromodiphenyl ether (BDE-47) in California women. Environ. Health Perspect. 111, 1175–1179. Pirard, C., De Pauw, E., Focant, J.F., 2003. Levels of selected PBDEs and PCBs in Belgian human milk. Organohalogen Compd. 64, 158–161. Roosens, L., Neels, H., Koppen, G., Schoeters, G., Nelen, V., van Larebeke, N., Blust, R., Covaci, A., 2007. PBDE levels in pooled serum samples of newborns, adolescents and adults from Flanders, Belgium. In: Proceedings of the 4th International Workshop on Brominated Flame Retardants, Amsterdam, The Netherlands, 24–27 April 2007.
175
Schecter, A., Papke, O., Joseph, J., Tung, K.C., 2005a. Polybrominated diphenyl ethers (PBDEs) in US computers and domestic carpet vacuuming: possible sources of human exposure. J. Toxicol. Environ. Health A 68, 501–513. Schecter, A., Papke, O., Tung, K.C., Joseph, J., Harris, T.R., Dahlgreen, J., 2005b. Polybrominated diphenyl ether flame retardants in the US population: Current levels, temporal trends, and comparison with dioxins, dibenzofurans, and polychlorinated biphenyls. J. Occup. Environ. Med. 47, 199–211. She, J.W., Petreas, M., Winkler, J., Visita, P., McKinney, M., Kopec, D., 2002. PBDEs in the San Francisco Bay area: measurements in harbor seal blubber and human breast adipose tissue. Chemosphere 46, 697–707. Sjödin, A., Patterson Jr., D.G., Bergman, Å., 2003. A review on human exposure to brominated flame retardants-particularly polybrominated diphenyl ethers. Environ. Int. 29, 829–839. Sjödin, A., Jones, R.S., Focant, J.F., Lapeza, C., Wang, R.Y., McGahee, M.E., Zhang, Y.L., Turner, W.E., Slazyk, B., Needham, L.L., Patterson Jr., D.G., 2004. Retrospective time-trend study of polybrominated diphenyl ether and polybrominated and polychlorinated biphenyl levels in human serum from the United States. Environ. Health Perspect. 112, 654–658. Smeds, A., Saukko, P., 2003. Brominated flame retardants and phenolic endocrine disrupters in Finnish human adipose tissue. Chemosphere 53, 1123–1130. Stapleton, H.M., Dodder, N.G., Offenberg, J.H., Schantz, M.M., Wise, S.A., 2005. Polybrominated diphenyl ethers in house dust and clothes dryer lint. Environ. Sci. Technol. 39, 925–931. Strandman, T., Koistinen, J., Kiviranta, H., Vuorinen, P.J., Tuomisto, J., Vartiainen, T., 1999. Levels of some polybrominated diphenyl ethers (PBDEs) in fish and human adipose tissue in Finland. Organohalogen Compd. 40, 355–358. Thomsen, C., Lundanes, E., Becher, G., 2002. Brominated flame retardants in archived serum samples from Norway: a study on temporal trends and the role of age. Environ. Sci. Technol. 36, 1414–1418. Van Wouwe, N., Covaci, A., Kannan, K., Gordon, J., Chu, A., Eppe, G., De Pauw, E., Goeyens, L., 2004. Levels of contamination for various pollutants present in Belgian human plasma. Organohalogen Compd. 66, 2818–2824. Voorspoels, S., Covaci, A., Maervoet, J., Schepens, P., 2002. Relationship between age and levels of organochlorine contaminants in human serum of a Belgian population. Bull. Environ. Contam. Toxicol. 69, 22–29. Voorspoels, S., Covaci, A., Neels, H., Schepens, P., 2007. Dietary PBDE intake: a market basket study in Belgium. Environ. Int. 33, 93–97. Voorspoels, S., Covaci, A., Schepens, P., 2003. Polybrominated diphenyl ethers in marine species from the Belgian North Sea and the Western Scheldt Estuary: Levels, profiles, and distribution. Environ. Sci. Technol. 37, 4348–4357. World Wide Found. Stockholm Convention ‘‘New POPs”: Screening additional POPs candidates. April 2005, 38 p. Wu, N., Herrmann, T., Paepke, O., Tickner, J., Hale, R., Harvey, E., La Guardia, M., McClean, M.D., Webster, T.F., 2007. Human exposure to PBDEs: associations of PBDE body burdens with food consumption and house dust concentrations. Environ. Sci. Technol. 41, 1584–1589.
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in human liver and adipose tissue samples from Belgium Adrian Covaci a,b,*, Stefan Voorspoels a, Laurence Roosens a, Werner Jacobs c, Ronny Blust b, Hugo Neels a a
Toxicological Centre, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk-Antwerp, Belgium Ecophysiology, Biochemistry and Toxicology Group, Department of Biology, University of Antwerp, Groenenborgerlaan 171, 2020 Antwerp, Belgium c Department of Forensic Medicine, University Hospital of Antwerp, Wilrijkstraat 10, 2650 Antwerp, Belgium b
a r t i c l e
i n f o
Article history: Received 2 December 2007 Received in revised form 23 February 2008 Accepted 26 February 2008 Available online 11 April 2008 Keywords: Polybrominated diphenyl ethers Polychlorinated biphenyls Human liver Human adipose tissue Age and gender dependency Belgium
a b s t r a c t Levels of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) were measured in paired human adipose tissue and liver samples (n = 25) from Belgium. Average concentrations and standard deviation of sum PBDEs (congeners 28, 47, 99, 100, 153, 154 and 183) were 5.3 ± 3.0 (range 1.4–13.2) and 3.6 ± 2.1 (range 1.0–10.0) ng g1 lipid weight (lw) in adipose tissue and liver, respectively. These concentrations were similar to reported PBDE data from Belgium and were at the lower end of the concentration range reported elsewhere in the world. In both tissues under study, BDE 153 and BDE 47 were the most abundant PBDE congeners, contributing approximately 35% and 25% to the total PBDE content. Average concentrations and range of PCBs (sum of 23 congeners) were 490 (range 70–1130) and 380 (range 90–1140) ng g1 lw in adipose tissue and liver, respectively. No correlation between age and concentrations of PBDEs could be found (r = 0.04), while PCB concentrations correlated significantly with age (r = 0.62, p < 0.01, for the sum PCBs; r = 0.64, p < 0.01 for PCB 153 alone). Factors, such as exposure pathways (food, dust and air), rates of bioaccumulation, metabolism and elimination, influence the concentrations of PBDEs differently than those of PCBs in humans. Ó 2008 Elsevier Ltd. All rights reserved.
1. Introduction The massive use of polybrominated diphenyl ethers (PBDEs) as flame retardants in thermoplastics (e.g. computer and TV housing), textiles, foams, interiors of cars, buses and airplanes is related to strict fire regulations. As a consequence, these products are extensively present in our daily life. Concurrent with the increasing use, environmental levels of PBDEs have risen since their first application. Spillage and emission during production and use, release from the consumer products when used and also disposal at the end-oflife of the consumer products, account for this phenomenon. These compounds are chemically and biologically persistent, lipophilic, bioaccumulate in fatty tissues and biomagnify throughout food chains (Law et al., 2006). Recently, the constituents of Penta-BDE technical mixtures (mainly consisting of tri- to hexa-BDE congeners) have been proposed as persistent organic pollutants (POPs) candidates since they
* Corresponding author. Address: Toxicological Centre, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk, Belgium. Tel.: +32 3 820 2704; fax: +32 3 820 2722. E-mail address: [email protected] (A. Covaci). 0045-6535/$ - see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2008.02.059
embody all characteristics of the Stockholm convention definition of POPs: bioaccumulation, toxicity, persistency and long-range transport potential (WWF, 2005). Penta-BDE mixtures, as well as the octa-BDE mixture, have been banned in 2004 in the European Union (EEC, 2003). Although the human exposure pathways to these chemicals through food and dust is increasingly being investigated (Schecter et al., 2005a; Stapleton et al., 2005; Harrad and Diamond, 2006a; Voorspoels et al., 2007; Harrad et al., 2008), any causality with the observed levels in human tissues is still far from understood. Recently, Wu et al. (2007) have reported a significant positive correlation between PBDEs in human milk and in dust from the homes of donors (n = 11), suggesting that dust might be an important exposure pathway for humans. The existing database with levels of PBDEs in humans is yet limited and contains mostly data from serum and milk (Thomsen et al., 2002; Sjödin et al., 2003, 2004; Gill et al., 2004; Schecter et al., 2005b). The present study aims to expand the knowledge concerning human exposure to PBDEs and PCBs in a Western European country and to discuss the extent of contamination and distribution of these contaminants two lesser studied tissues, namely liver and adipose tissue. Additionally, we explored the relationships between age or gender and levels of PBDEs and PCBs.
171
A. Covaci et al. / Chemosphere 73 (2008) 170–175
2. Materials and methods
2.3. Quality assurance
2.1. Sample collection
Retention times, ion chromatograms and relative abundance of the monitored ions were used as identification criteria. A deviation of ion abundance ratios within 15% of the mean values for calibration standards was considered acceptable. Quantification was based on five-point calibration curves in which either the sum of m/z = 79 and 81 for PBDEs or the most intense ion from the molecular cluster for PCBs were monitored. The method performance was assessed through rigorous internal quality control, which included daily check of calibration curves, regular analysis of method blanks and certified reference materials (BCR 350 – PCBs in mackerel oil and SRM 1945 – PCBs and PBDEs in whale blubber). For quality assurance and control, the laboratory participated in the Interlaboratory Comparison Exercise Program for Organic Contaminants in Marine Mammal Tissues organized by the National Institute of Standards and Technology (NIST, Gaithersburg, MD, USA). Results for individual PCB and PBDE congeners had a variation coefficient less than 15% from the target values (Kucklick et al., 2005, 2007). Since BDE 153 was found in unusual high proportion in several samples, special attention was given to its quality assurance. As previously reported by us (Korytar et al., 2005), BDE 153 was not co-eluted with other major PBDE congeners or with hexabromocyclododecane. Spiking experiments (n = 3) at 5 ng/g lw done in pork liver indicated a recovery of 86 ± 4%. Further, results for BDE 153 (together with BDE 47 and CB 153) obtained during participation in different interlaboratory exercises are given in Table 1. Some degradation probably arising from both the column and injector was observed for PBDE congeners with 8–10 bromine atoms and therefore, no further efforts were made to quantify these congeners, which might have been present in the samples. For PBDE and PCB congeners that were consistently measured in the procedural blanks, the mean blank value was subtracted. BDE 47 and BDE 99 were the only PBDE congeners detected in the blanks with mean value (SD) of 0.07 ± 0.03 ng and 0.06 ± 0.02 ng, respectively. The limit of quantification (LOQ) was calculated as 2 SD of the blank values and related to the sample intake, namely 0.2 and 5 g for adipose tissue and liver, respectively (Voorspoels et al., 2003). For PBDE (including BDE 153) and PCB congeners not detected in the blanks, the LOQ was calculated based on a instrumental signal-to-noise of 10. LOQs ranged between 0.03 and 0.2 ng g1 lipid weight (lw) for tri- to hepta-BDEs and between 1 and 4 ng g1 lw for tri- to octa-CBs.
The present study was approved by the Ethics Committees of the University Hospital of Antwerp and of the University of Antwerp. Paired human liver and adipose tissue (mesenteric fat) samples (n = 25) were collected between 2003 and 2005 during autopsy at the Department of Forensic Medicine of the University Hospital of Antwerp. Samples originated from 18 males and 7 females, who died from causes unrelated to environmental contaminants as far as could be determined. The mean age of the individuals was 37 years (range 9–70 years), while the mean weight was 75 kg (range 30–100 kg). No information about possible occupational exposure or dietary habits was available. Samples were collected in hexane-washed polyethylene recipients, frozen immediately and stored at 20 °C until analysis. 2.2. Chemical analysis The sample preparation and analysis method was previously validated and reported elsewhere in detail (Covaci et al., 2002; Voorspoels et al., 2003). Based on their reported abundance, the following contaminants (IUPAC numbering), were targeted for analysis: PBDE congeners (Nos. 28, 47, 99, 100, 153, 154 (co-eluting with BB 153), and 183) and PCB congeners (Nos. 28/31, 52, 74, 95, 99, 101, 105, 110, 118, 128, 138/163, 149, 153, 156, 170, 180, 183, 187, 194, 196, and 196). The individual PCB standards were obtained from Dr. Ehrenstorfer Laboratories (Augsburg, Germany) and PBDE standard mixtures were obtained from Wellington Laboratories (Guelph, Ontario, Canada). Hexane, acetone, dichloromethane, iso-octane (all pesticide grade) and concentrated sulfuric acid (analytical grade) were obtained from Merck (Darmstadt, Germany). Silica gel 60 (63–230 mesh) and anhydrous Na2SO4 (Merck) were heated at 150 °C for 24 h. Liver (3 g) or adipose tissue samples (0.3 g) were mixed with anhydrous Na2SO4, spiked with internal standards (BDE 77, BDE 128, CB 46 and CB 143) and Soxhlet-extracted with hexane:acetone (3:1, v/v). After gravimetrical lipid determination (on an aliquot), the extract was cleaned on acidified silica gel and the analytes were eluted with 15 ml n-hexane and 10 ml dichloromethane. The eluate was concentrated to near dryness and reconstituted in 100 ll iso-octane. The analyses were performed on an Agilent 6890 GC coupled with an Agilent 5973 MS. For PCBs, electron ionization (EI) mode was used in combination with a 30 m 0.25 mm 0.25 lm DB-1 capillary column (J&W Scientific, Folsom, CA, USA). For PBDEs, electron capture negative ionization (ECNI) mode was used in combination with a 12 m 0.18 mm 0.20 lm AT-5 column (Alltech, Lokeren, Belgium). Instrumental parameters were presented in detail by Covaci et al. (2002) and Voorspoels et al. (2003).
2.4. Data treatment Concentrations below the LOQ were assigned a value of f*LOQ, with ‘f ’ the proportion of measurements with levels above the LOQ (Voorspoels et al., 2002). This approach allowed data below LOQ to be used in the statistical data treatment. All statistical
Table 1 Values obtained by the Toxicological Center for BDE 47, BDE 153 and CB 153 in recent interlaboratory exercises BDE 47
AMAP 2007 Round 1 (ng/ml)
Round 3 (ng/ml)
NIST 2005 Unknown blubber (ng/g) SRM 1945 (ng/g)
BDE 153
Mean ± SD
Assigned
0.36 0.90 0.99 0.74 0.44 0.99
0.44 0.97 0.99 0.73 0.44 0.98
123 ± 4 36.4 ± 0.5
132 ± 17 39.6 ± 0.2
Mean ± SD 0.19 0.37 0.24 0.33 0.21 0.44 12.6 ± 0.4 7.5 ± 0.5
CB 153 Assigned 0.22 0.42 0.31 0.35 0.25 0.46 14.2 ± 1.6 8.3 ± 0.6
Mean ± SD
Assigned
1.93 3.47 1.17 3.12 3.91 1.09
1.9 3.1 1.2 3.0 3.9 1.1
1700 ± 40 211 ± 3
1830 ± 240 228 ± 10
172
A. Covaci et al. / Chemosphere 73 (2008) 170–175
analyses were performed using Statistica for Windows and GraphPad Instat version 3.06 for Windows. The normal distribution of concentrations of each congener within the dataset for tissue was evaluated using both Shapiro-Wilks and K–S tests. Since concentrations were normally distributed, a paired t-test was used to test differences between PBDE concentrations in liver and adipose tissue. No outliers were detected using Grubbs’ test. Pearson correlation coefficients were used to measure the strength of the association between age and the organohalogen concentrations. 3. Results and discussion 3.1. Residue levels of PBDEs and comparison with other data The average lipid percentage and standard deviation in liver was 7.0 ± 3.8% (range 3.3–16.9%), while the average lipid percentage in adipose tissue was 86.2 ± 6.1% (range 70.1–94.1%). All investigated PBDE congeners had a detection frequency higher than 80%, with BDE 153 and BDE 154/BB 153 being detected in all samples. The co-elution between BDE 154 and BB 153 could not be resolved on the AT-5 column used with the GC-ECNI/MS detection (Covaci et al., 2003, 2007). However, since BDE 154 is reported as a minor PBDE congener in humans (Covaci et al., 2002; Sjödin et al., 2003), we can assume that BB 153 had in these samples a higher contribution to the peak. This assumption is also strengthened by the fact that BB 153 was previously measured in serum of Belgian donors, while BDE 154 could not be detected in the same samples (Covaci and Voorspoels, 2005). Average concentration of sum PBDEs (sum of congeners 28, 47, 99, 100, 153) in Belgian liver samples was 3.6 ± 2.1 ng g1 lw, range 1.0–10.0 ng g1 lw (Table 2). For adipose tissue, average of sum PBDEs was 5.3 ± 3.0 ng g1 lw, range 1.4–13.2 ng g1 lw (Table 2). There were no outliers between the concentrations of PBDEs in the present study, while other studies have reported the presence of 5% values with high concentrations (Harrad and Porter, 2007). PBDE concentrations in the present study were similar to other reported concentrations in Belgium. Table 3 summarizes all available Belgian data in different human matrices (adipose tissue, milk and serum). Concentrations of PBDEs in Belgian adipose tissue and liver were similar to reported concentrations in tissues from other European countries (Strandman et al., 1999; Smeds and Saukko, 2003; Fernandez et al., 2007; Meneses et al., 1999; Meironyté Guvenius et al., 2001) and were notably lower than those in USA (Petreas et al., 2003; Johnson-Restrepo et al., 2005). In a Swedish survey, BDE 47 was measured in more than 400 individuals (Hardell
Table 2 Average concentrations (ng g1 lw), standard deviations, detection frequency and range (in brackets) of PBDE congeners in paired liver and adipose tissue samples from 25 Belgian individuals Detection frequency (%)
Adipose tissue (ng g1 lw)
Liver (ng g1 lw)
84 96 86 94 100 100 96
25 86 ± 6.1% (70–93%) 0.08 ± 0.06a 1.2 ± 1.1a 0.55 ± 0.47a 0.34 ± 0.33b 2.0 ± 1.8b 0.91 ± 0.59b 0.31 ± 0.12b
25 7.0 ± 3.8% (3.1–16.9%) 0.06 ± 0.04a 0.95 ± 0.83a 0.38 ± 0.36a 0.17 ± 0.20b 1.2 ± 1.4b 0.66 ± 0.43b 0.21 ± 0.12b
96 100 100 100 100 100 100 100
5.3 ± 3.0b (1.4–13.2) 9.7 ± 6.4a 21.0 ± 18.1a 83.0 ± 58.5a 131 ± 92b 40.5 ± 29.4a 98.1 ± 74.4a 10.5 ± 7.4b 26.1 ± 21.5b
3.6 ± 2.1b (1.0–10.0) 8.1 ± 8.0a 19.1 ± 18.1a 68.2 ± 55.0a 99.6 ± 80.4b 33.6 ± 24.3a 70.7 ± 57.8a 8.4 ± 6.8b 22.4 ± 19.0b
Sum 7 marker PCBs
334 ± 234b (47–798)
259 ± 205b (57–756)
Sum PCBs
490 ± 341b (68–1128)
377 ± 299b (89–1138)
N Lipids BDE BDE BDE BDE BDE BDE BDE
28 47 99 100 153 154/BB 153 183
Sum PBDEs CB CB CB CB CB CB CB CB
a b
99 118 138 + 163 153 170 180 183 187
Not significantly different. Significantly different (p < 0.05).
et al., 1998). Concentrations ranged between 0.5 and 4 ng g1 lw for most specimens with few persons having levels up to 100 ng g1 lw, indicating a large variation in the individual values. Adipose tissue samples from USA (Petreas et al., 2003) contained almost one order of magnitude higher concentrations of BDE 47 than our group, while even higher concentrations were measured in a newer study (Johnson-Restrepo et al., 2005). Concentrations of PBDEs in Japanese individuals were similar to or even lower than the Belgian concentrations (Choi et al., 2007; Kunisue et al., 2007). The differences between countries can be partially explained by the PBDE concentrations in the diet and especially in food items with higher contribution to the total PBDE intake (fish and meat products) (Voorspoels et al., 2007). However, other exposure pathways (e.g. indoor dust) have recently been identified for humans and their contribution to the total PBDE intake may be higher than
Table 3 Summary of PBDE concentrations (ng g1 lw) in Belgian human samples Sampling year
Sample type
Number of samples
Age
Sum PBDEsa (ng g1 lw) mean ± SD
% BDE 47 and BDE 153 to sum PBDEsa
Reference
1999–2004
Serum
11
n.a.
4.0 ± 1.6
31 and 43
2000–2001 2001 2002–2003
Milk Adipose tissue Pooled cord serum Pooled cord serum Serum Pooled serum Pooled serum Adipose tissue Liver Adipose tissue
14 20 4
26–38 19–77 0
2.9 4.7 ± 2.3 2.6 ± 0.4
59 and 15 31 and 52 49 and 25
0
2.2 ± 0.5
19 and 40
Covaci and Voorspoels 2005 Pirard et al., 2003 Covaci et al., 2002 Covaci and Voorspoels 2005 Roosens et al., 2007
19–63 14–15 50–65
3.4 ± 3.4 4.0 ± 0.4 4.6 ± 0.8 7.6 ± 8.9 3.6 ± 2.1 5.3 ± 3.0
36 32 24 22 26 22
Van Wouwe et al., 2004 Roosens et al., 2007 Roosens et al., 2007 Naert et al., 2006 This study This study
2002–2003 2003 2003–2004 2004–2005 2006 2002 and 2006 2002 and 2006
7 20 8 8 53 25 25
9–70 9–70
n.a. – Not available. a For each study, sum PBDEs = BDE 28 + 47 + 99 + 100 + 153 + 154 + 183.
and and and and and and
35 36 30 49 33 37
173
A. Covaci et al. / Chemosphere 73 (2008) 170–175
expected, at least for some groups, such as toddlers (Harrad et al., 2008). 3.2. Congener profile of PBDEs BDE 153 and BDE 47 were the most abundant PBDE congeners in liver, attributing for 33% and 26%, respectively. Similar to the liver, these two congeners were also the most abundant in adipose tissue, contributing to 37% and 22% to the total PBDE content (Fig. 1). High concentrations of BDE 153 were also seen in some samples from Spain (Meneses et al., 1999). BDE 153, as the dominant congener, was also observed in Japan (Choi et al., 2007; Kunisue et al., 2007), Spain (Meneses et al., 1999; Fernandez et al., 2007), Sweden (Meironyté Guvenius et al., 2001) and Belgium (Covaci et al., 2002). However, in the US, where greater amounts of technical Penta-BDE have been used, BDE 47 was by far the most predominant congener (60–70% of the total PBDEs) in human adipose tissue (Johnson-Restrepo et al., 2005; She et al., 2002). The difference in the profiles of PBDEs might be related to dietary habits or to temporal trends in the exposure to various PBDE technical mixtures. Furthermore, exposure through dust can have an important role in the resulting PBDE profiles present in the exposed individuals. The more lipophilic PBDE congeners (BDEs 100, 153, and 183) and the sum PBDEs had slightly higher statistically significant concentrations in adipose tissue compared to liver samples (Table 2 and Fig. 1). This can be explained by the relatively low metabolic activity of adipose tissue which is primarily used for energy (and pollutants) storage, while liver is an active tissue where contaminants are (eventually) more readily metabolized. Furthermore, the profiles of PBDE congeners in adipose tissue or liver differed from that in human milk (Meironyté et al, 1999) and in blood (Thomsen et al., 2002), with BDE 47 being by far the predominant PBDE congener (70% of the total sum PBDEs). Obviously, the PBDE profile in human tissues depends on the biotransformation and accumulation kinetics of each congener. Furthermore, the distribution of PBDE congeners between different tissues could be directly related to their half-lives, octanol-water partition coefficient, and the properties of the tissue (Inoue et al. 2006). The high percentage of BDE 153 observed in the investigated adipose tissue and liver samples might also be a consequence of its long half-life (Sjödin et al., 2003), and hence it is likely that
BDE 153 is more persistent and lipid-dependently accumulated in human bodies. The concentrations of individual PBDE congeners and sum PBDEs in liver and adipose tissue were significantly correlated (BDE 153: r = 0.93, p < 0.01, CA = 1.21CL + 0.52; BDE 47: r = 0.80, p < 0.01, CA = 1.04CL + 0.20; sum PBDEs: r = 0.74, p < 0.01, CA = 1.07CL + 1.46), CA and CL are the concentrations in adipose tissue and liver, respectively. The lower correlation coefficients obtained for sum PBDEs may be explained by the higher number of no-detects and lower values for PBDE congeners other than BDE 47 and 153, which greatly contributed to the sum PBDEs. 3.3. Residues levels and congener profiles of PCBs Several PCB congeners (CB 31/28, CB 52 and CB 128) were not detected in any sample, while other PCBs (CB 74, CB 95, CB 105, CB 149) had a detection frequency of less than 50%. These congeners were therefore not included in the statistical analysis. As expected, the persistent PCB congeners (CB 99, CB 118, CB 138/163, CB 153, CB 170, CB 183, CB 187 and CB 183) were detected in all samples and their descriptive parameters are given in Table 2. Average concentration of sum PCBs (sum of 23 PCB congeners in Belgian liver samples was 377 ± 299 ng g1 lw, range 89– 1138 ng g1 lw (Table 2). For adipose tissue, average of sum PCBs was 490 ± 341 ng g1 lw, range 68–1128 ng g1 lw (Table 2). Concentrations of PCBs measured in the present study were similar to previous studies in Belgian adipose tissue (Covaci et al. 2002; Naert et al. 2006). Concentrations of the sum 7 marker PCBs are also given as additional information in Table 2. As previously reported (Covaci et al., 2002; Naert et al., 2006), CB 153 and CB 180 were the major PCB congeners in human adipose tissue, similar to profiles reported from current European reports. Only few congeners (CB 153, CB 183 and CB 187) together with sum PCBs presented higher statistically significant concentrations in adipose tissue than in liver. Similar to PBDEs, this is probably due to the storage role of adipose tissue compared to the more active function of the liver. The concentrations of selected individual PCB congeners and sum PCBs in liver and adipose tissue were significantly correlated (CB 153: r = 0.72, p < 0.01, CA = 0.83CL + 49; CB 180: r = 0.62, p < 0.05, CA = 0.80CL + 42; sum PCBs: r = 0.69, p < 0.01, CA = 0.78CL + 195).
45
40 liver adipose tissue
35
% of sum PBDEs
30
25
20
15
10
5
0 BDE 28
BDE 47
BDE 100
BDE 99
BDE 154/BB153
BDE 153 *
BDE 183
Fig. 1. Congener profiles of PBDEs in adipose tissue and liver. Error bars represent 2 standard error (SE).
174
A. Covaci et al. / Chemosphere 73 (2008) 170–175
3.4. Age- and gender-dependent accumulation of PCBs and PBDEs Although it was suggested that the accumulation of PBDEs will increase with age (Hooper and McDonald, 2000), no clear evidence could be found in the scientific literature. In the present study, PBDE concentrations in human tissues (data from Covaci et al 2002; Van Wouwe et al., 2004; Naert et al., 2006; Roosens et al., 2007) could not be correlated with age (r = 0.004 for sum PBDEs; r = 0.04 for BDE 153). It suggests that human exposure to PBDEs is of recent nature and that these contaminants have not yet attained a steady-state in human tissues. In contrast, PCBs correlated significantly with age (r = 0.62, p < 0.01, for the sum PCBs; r = 0.64, p < 0.01 for PCB 153) (Fig. 2). Interestingly, some of the highest concentrations of PBDEs were measured in young specimens. This suggests that other sources than long-term dietary exposure (e.g. indoor dust or air) are also important for PBDEs. However, several additional factors, such as the rate of metabolism of PBDE congeners, the type of technical mixture to which the persons have been exposed, mother–child transfer, or life-stage specific exposures (e.g. toddlers), should also be taken into account when assessing the accumulation of PBDEs in humans. Similar to our results, Hardell et al. (1998) did not find an age-dependency for levels of BDE 47 in adipose tissue from 77 individuals (age range 28–85 years) with different types of cancer. They also concluded that the determinant factors for the concentrations of PBDEs in humans are most probably different than those for PCBs and other POPs. Although the number of samples for each gender group in the present study was limited, a tendency for higher concentrations of sum PBDEs in males was observed in adipose tissue, but not in liver (t-test, p < 0.05). This result is consistent with observations of Kunisue et al. (2007), which have measured higher concentra-
a conc PBDEs (ng/g lw)
45 40
r = 0.004
35 30 25 20 15 10 5 0 0
20
40
60
80
100
age
b conc sum PCBs (ng/g lw)
2500
r = 0.62, p < 0.01 2000 1500 1000 500 0 0
20
40
60
80
100
age Fig. 2. Pearson correlation between (a) age and sum PBDEs and (b) between age and sum PCBs in Belgian human samples. PBDE data are from Van Wouwe et al. (2004), Covaci and Voorspoels (2005), Naert et al. 2006 and from the present study.
tions of PBDEs in Japanese adipose tissue from males. Therefore, breast-feeding might play an important role in the elimination of PBDEs in females. The higher concentrations in males may also be attributable to the fact that their lifestyles may involve spending more time in contaminated microenvironments, such as offices and cars, which have been shown to be more contaminated than homes (Harrad et al, 2006b). In another study, no age or gender dependency with PBDE concentrations has been recently seen in individuals from New Zealand (Harrad and Porter, 2007). Contrarily, no gender differences for PCBs were observed in liver or in adipose tissue, which is also consistent with the findings of Kunisue et al. (2007). This indicates that the recent human exposure to PCBs is relatively constant, while the concentrations of PCBs in human tissues have already reached a steady-state. Acknowledgements Katrien Jorissen is thanked for the sample preparation. Dr. Caroline Naert is acknowledged for providing individual data on PCBs, PBDEs and age from her study (Naert et al., 2006). Adrian Covaci acknowledges a postdoctoral fellowship of the Funds for Scientific Research Flanders (FWO). The authors thank to the organizers of the BFR 2007 workshop in Amsterdam for allowing this study to be presented as a poster. References Choi, J., Fujimaki, S., Kitamura, K., Hashimoto, S., Ito, H., Suzuki, N., Sakai, S., Morita, M., 2007. Polybrominated dibenzo-p-dioxins, dibenzofurans, and diphenyl ethers in Japanese human adipose tissue. Environ. Sci. Technol. 37, 817–821. Covaci, A., de Boer, J., Ryan, J.J., Voorspoels, S., Schepens, P., 2002. Distribution of organobrominated and organochlorinated contaminants in Belgian human adipose tissue. Environ. Res. 88, 210–218. Covaci, A., Voorspoels, S., de Boer, J., 2003. Determination of BFRs with emphasis on PBDEs in environmental and human samples – a review. Environ. Int. 29, 735– 756. Covaci, A., Voorspoels, S., 2005. Optimization of the determination of polybrominated diphenyl ethers in human serum using solid-phase extraction and gas chromatography-electron capture negative ionization mass spectrometry. J. Chromatogr. B 827, 216–223. Covaci, A., Voorspoels, S., Ramos, L., Neels, H., Blust, R., 2007. Recent developments in the analysis of brominated flame retardants and brominated natural compounds. J. Chromatogr. A 1153, 145–171. EEC, 2003. Directive 2003/11/EC of the European Parliament and of the Council of 6 February 2003 amending for the 24th time Council directive 76/769/EEC relating to restrictions on the marketing and use of certain dangerous substances and preparations pentabromodiphenyl ether and octabromodiphenyl ether). Official J. L 042, 15/02/2003. Fernandez, M.F., Araque, P., Kiviranta, H., Molina-Molina, J.M., Rantakokko, P., Laine, O., Vartiainen, T., Olea, N., 2007. PBDEs and PBBs in the adipose tissue of women from Spain. Chemosphere 66, 377–383. Gill, U., Chu, I., Ryan, J.J., Feeley, M., 2004. Polybrominated diphenyl ethers: human tissue levels and toxicology. Rev. Environ. Contam. Toxicol. 183, 55–97. Hardell, L., Lindström, G., van Bavel, B., Wingfors, H., Sundelin, E., Liljegren, G., 1998. Concentrations of the flame retardant 2,20 ,4,40 -tetrabrominated diphenyl ether (BDE 47) in human adipose tissue in Swedish persons and the risk for nonHodgkin’s lymphoma. Oncol. Res. 10, 429–432. Harrad, S., Diamond, M., 2006a. New directions: exposure to polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs): current and future scenarios. Atmos. Environ. 40, 1187–1188. Harrad, S., Hazrati, S., Ibarra, C., 2006b. Concentrations of polybrominated diphenyl ethers in indoor air and dust and polychlorinated biphenyls in indoor air in Birmingham, United Kingdom: implications for human exposure. Environ. Sci. Technol. 40, 4633–4638. Harrad, S., Porter, L., 2007. Concentrations of polybrominated diphenyl ethers in blood serum from New Zealand. Chemosphere 66, 2019–2023. Harrad, S., Ibarra, C., Diamond, M., Melymuk, L., Robson, M., Douwes, J., Roosens, L., Dirtu, A.C., Covaci, A., 2008. Polybrominated diphenyl ethers in domestic indoor dust from Canada, New Zealand, United Kingdom and United States. Environ. Int. 34, 232–238. Hooper, K., McDonald, T.A., 2000. The PBDEs: an emerging environmental challenge and another reason for breast-milk monitoring programs. Environ. Health Perspect. 108, 387–392. Inoue, K., Harada, K., Takenaka, K., Uehara, S., Kono, M., Shimizu, T., Takasuga, T., Senthilkumar, K., Yamashita, F., Koizumi, A., 2006. Levels and concentration ratios of polychlorinated biphenyls and polybrominated diphenyl ethers in serum and breast milk in Japanese mothers. Environ. Health Perspect. 114, 1179–1185.
A. Covaci et al. / Chemosphere 73 (2008) 170–175 Johnson-Restrepo, B., Kannan, K., Rapaport, D.P., Rodan, B.D., 2005. Polybrominated diphenyl ethers and polychlorinated biphenyls in human adipose tissue from New York. Environ. Sci. Technol. 39, 5177–5182. Korytar, P., Covaci, A., de Boer, J., Gelbin, A., Brinkman, U.A.Th., 2005. Retention-time database of 126 PBDE congeners and 2 Bromkal technical mixtures on 7 capillary GC columns. J. Chromatogr. A 1065, 239–249. Kucklick, J.R., Pugh, R.S., Becker, P.R., Schantz, M.M., Wise, S.A., Rowles, T.K., 2005. Description and results of the 2003 NIST/NOAA interlaboratory comparison exercise program for organic contaminants in marine mammal tissues. NISTIR 7269, 157 p. Kucklick, J.R., Pugh, R.S., Becker, P.R., Schantz, M.M., Wise, S.A., Rowles, T.K. Description and results of the 2005 NIST/NOAA interlaboratory comparison exercise program for organic contaminants in marine mammal tissues. NISTIR 7410, 132 p. Kunisue, T., Takayanagi, N., Isobe, T., Takahashi, S., Nose, M., Yamada, T., Komori, H., Arita, N., Ueda, N., Tanabe, S., 2007. Polybrominated diphenyl ethers and persistent organochlorines in Japanese human adipose tissues. Environ. Int. 33, 1048–1056. Law, R.J., Allchin, C.R., de Boer, J., Covaci, A., Herzke, D., Lepom, P., Morris, S., Tronczynski, J., de Wit, C.A., 2006. Levels and trends of brominated flame retardants in the European environment. Chemosphere 64, 187– 208. Meneses, M., Wingfors, H., Schuhmacher, M., Domingo, J.L., Lindström, G., van Bavel, B., 1999. Polybrominated diphenyl ethers detected in human adipose tissue from Spain. Chemosphere 39, 2271–2278. Meironyté, D., Norén, K., Bergman, Å., 1999. Analysis of polybrominated diphenyl ethers in Swedish human milk. A time-related trend study, 1972–1997. J. Toxicol. Environ. Health 58, 329–341. Meironyté Guvenius, D., Bergman, Å., Norén, K., 2001. Polybrominated diphenyl ethers in Swedish human liver and adipose tissue. Arch. Environ. Contam. Toxicol. 40, 564–570. Naert, C., Piette, M., Bruneel, N., Van Peteghem, C., 2006. Occurrence of polychlorinated biphenyls and polybrominated diphenyl ethers in Belgian human adipose tissue samples. Arch. Environ. Contam. Toxicol. 50, 290–296. Petreas, M., She, J., Brown, F.R., Winkler, J., Windham, G., Rogers, E., Zhao, G.M., Bhatia, R., Charles, M.J., 2003. High body burdens of 2,20 ,4,40 tetrabromodiphenyl ether (BDE-47) in California women. Environ. Health Perspect. 111, 1175–1179. Pirard, C., De Pauw, E., Focant, J.F., 2003. Levels of selected PBDEs and PCBs in Belgian human milk. Organohalogen Compd. 64, 158–161. Roosens, L., Neels, H., Koppen, G., Schoeters, G., Nelen, V., van Larebeke, N., Blust, R., Covaci, A., 2007. PBDE levels in pooled serum samples of newborns, adolescents and adults from Flanders, Belgium. In: Proceedings of the 4th International Workshop on Brominated Flame Retardants, Amsterdam, The Netherlands, 24–27 April 2007.
175
Schecter, A., Papke, O., Joseph, J., Tung, K.C., 2005a. Polybrominated diphenyl ethers (PBDEs) in US computers and domestic carpet vacuuming: possible sources of human exposure. J. Toxicol. Environ. Health A 68, 501–513. Schecter, A., Papke, O., Tung, K.C., Joseph, J., Harris, T.R., Dahlgreen, J., 2005b. Polybrominated diphenyl ether flame retardants in the US population: Current levels, temporal trends, and comparison with dioxins, dibenzofurans, and polychlorinated biphenyls. J. Occup. Environ. Med. 47, 199–211. She, J.W., Petreas, M., Winkler, J., Visita, P., McKinney, M., Kopec, D., 2002. PBDEs in the San Francisco Bay area: measurements in harbor seal blubber and human breast adipose tissue. Chemosphere 46, 697–707. Sjödin, A., Patterson Jr., D.G., Bergman, Å., 2003. A review on human exposure to brominated flame retardants-particularly polybrominated diphenyl ethers. Environ. Int. 29, 829–839. Sjödin, A., Jones, R.S., Focant, J.F., Lapeza, C., Wang, R.Y., McGahee, M.E., Zhang, Y.L., Turner, W.E., Slazyk, B., Needham, L.L., Patterson Jr., D.G., 2004. Retrospective time-trend study of polybrominated diphenyl ether and polybrominated and polychlorinated biphenyl levels in human serum from the United States. Environ. Health Perspect. 112, 654–658. Smeds, A., Saukko, P., 2003. Brominated flame retardants and phenolic endocrine disrupters in Finnish human adipose tissue. Chemosphere 53, 1123–1130. Stapleton, H.M., Dodder, N.G., Offenberg, J.H., Schantz, M.M., Wise, S.A., 2005. Polybrominated diphenyl ethers in house dust and clothes dryer lint. Environ. Sci. Technol. 39, 925–931. Strandman, T., Koistinen, J., Kiviranta, H., Vuorinen, P.J., Tuomisto, J., Vartiainen, T., 1999. Levels of some polybrominated diphenyl ethers (PBDEs) in fish and human adipose tissue in Finland. Organohalogen Compd. 40, 355–358. Thomsen, C., Lundanes, E., Becher, G., 2002. Brominated flame retardants in archived serum samples from Norway: a study on temporal trends and the role of age. Environ. Sci. Technol. 36, 1414–1418. Van Wouwe, N., Covaci, A., Kannan, K., Gordon, J., Chu, A., Eppe, G., De Pauw, E., Goeyens, L., 2004. Levels of contamination for various pollutants present in Belgian human plasma. Organohalogen Compd. 66, 2818–2824. Voorspoels, S., Covaci, A., Maervoet, J., Schepens, P., 2002. Relationship between age and levels of organochlorine contaminants in human serum of a Belgian population. Bull. Environ. Contam. Toxicol. 69, 22–29. Voorspoels, S., Covaci, A., Neels, H., Schepens, P., 2007. Dietary PBDE intake: a market basket study in Belgium. Environ. Int. 33, 93–97. Voorspoels, S., Covaci, A., Schepens, P., 2003. Polybrominated diphenyl ethers in marine species from the Belgian North Sea and the Western Scheldt Estuary: Levels, profiles, and distribution. Environ. Sci. Technol. 37, 4348–4357. World Wide Found. Stockholm Convention ‘‘New POPs”: Screening additional POPs candidates. April 2005, 38 p. Wu, N., Herrmann, T., Paepke, O., Tickner, J., Hale, R., Harvey, E., La Guardia, M., McClean, M.D., Webster, T.F., 2007. Human exposure to PBDEs: associations of PBDE body burdens with food consumption and house dust concentrations. Environ. Sci. Technol. 41, 1584–1589.