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Inverse age-dependent accumulation of decabromodiphenyl ether and
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other PBDEs in serum from a general adult population
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Mercè Garí and Joan O. Grimalt *
Department of Environmental Chemistry, Institute of Environmental Assessment and Water
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Research (IDAEA). Spanish Council for Scientific Research (CSIC). Jordi Girona, 18. 08034
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Barcelona. Catalonia. Spain
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*
Author for correspondence. Phone: +34 93 4006118; Fax: +34 93 2045904; E-mail:
joan.grimalt@idaea.csic.es
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Keywords
Age-dependence; Catalonia; Decabromodiphenyl ether; Human exposure; Polybromodiphenyl
ethers; Serum samples
Abbreviations: AMAP, Arctic Monitoring and Assessment Programme; BDE, brominated diphenyl ether; BMI, body
mass index; LD, limit of detection; LQ, limit of quantification; MS, mass spectrometry; NHANES, National Health and
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Nutrition Examination Survey; NICI, negative ion chemical ionization; PBDE, polybromodiphenyl ether; POP,
persistent organic pollutant.
Abstract
Polybromodiphenyl ethers (PBDEs), including the decabromodiphenyl congener (BDE-209), were
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determined in serum of 731 individuals from a general adult population (18-74 years) collected in
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2002 in Catalonia (north-eastern Spain). The BDE-209 was the predominant congener (median 3.7
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ng/g lipid) followed by BDE-47 (2.6 ng/g lipid) and BDE-99 (1.2 ng/g lipid). PBDEs in this
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population (median 15.4 ng/g lipid) ranked among the highest of previously described
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concentrations in populations in Europe, Asia, New Zealand and Australia, yet it was lower than
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those found in North American reports. Age was clearly the socio-demographic factor of highest
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influence on the PBDE distributions. However, unlike usual trends of higher accumulation of POPs
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through age, the higher concentrations were found in young individuals (< 30 years) rather than in
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adults (≥ 30 years), with differences of 14%, 31% and 46% in the most abundant congeners (i.e.
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BDE-209, BDE99 and BDE-47, respectively). This age-dependent distribution of PBDEs (including
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the case for BDE-209, which is shown for the first time in this study) is explained by the higher and
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widespread use of these compounds since the 1980s. In view that these compounds remain highly
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used, this accumulation pattern is likely to evolve, anticipating an increasing level of PBDE
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concentrations in future general population surveys, yet probably assuming an age-dependent
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increase pattern. Socio-economic level was also a determinant of BDE-47 concentrations, but only
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relevant for the least affluent class, suggesting that lifestyle and environmental conditions in the
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dwelling place may also contribute to exposure. Nonetheless, gender, body mass index, place of
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birth, parity and education level did not show any statistically significant influence on the observed
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PBDE distributions.
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1. Introduction
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Polybromodiphenyl ethers (PBDEs) have been used as flame retardants for some decades in a wide
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range of products. They have been distributed in three commercial mixtures of congeners with
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different levels of bromination (penta-BDE, octa-BDE, and deca-BDE), which are named after the
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dominating homologue group. The pentabrominated formulation was the major source of BDE-47
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and BDE-99 congeners and was mainly employed as additive of polyurethane foams in furniture,
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carpets and bedding. The octabrominated mixture was dominated by BDE-183 followed by BDE-
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153 and BDE-154, being used in flame-retardant thermoplastics, such as high impact polystyrene.
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The decabrominated product was essentially composed of decabromodiphenyl ether (BDE-209) and
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was used predominantly for textiles and plastics for a variety of electronic products, in particular
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TVs and computers. According to the PBDE global market demands in 2001, the deca-BDE
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formulation was the dominant one (83%), followed by penta-BDE (11%) and octa-BDE (6%)
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(Guardia et al., 2006).
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PBDEs are semi-volatile and persistent in the environment. They accumulate in lipids and
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biomagnify through the food web (Johnson-Restrepo et al., 2005). They have been found in various
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environmental compartments as well as in human fluids and tissues (Hites, 2004). Total PBDE
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(ΣPBDE) concentrations in human serum have increased in the past decades, as reported in time
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trend studies conducted in the US (Sjödin et al., 2004) and Europe (Schröter-Kermani et al., 2000;
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Thomsen et al., 2002), including increases from 0.44 ng/g lipid (1977) to 3.3 ng/g lipid (1999) in
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Norway (Thomsen et al., 2002). However, as PBDEs are relatively new in the environment, their
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concentrations are still lower than those of other environmental pollutants of similar chemical
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properties, such as polychlorobiphenyls (Guvenius et al., 2003), that have been present for a longer
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period. Increasing concentrations but still lower than other persistent organic pollutants used in the
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past raise the need for impact assessment studies to examine trends and anticipate future deleterious
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effects.
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Human exposure to PBDEs may occur in many daily life situations. Diet is one of the main
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sources, e.g. fatty fish (Sjödin et al., 2000) or red meat and poultry (Fraser et al., 2009). Several
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studies reporting the concentrations of PBDEs in a number of foodstuffs are available (Bocio et al.,
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2003; Domingo et al., 2006; Gómara et al., 2006; Schecter et al., 2005). The indoor environment
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and dust ingestion, both at home and in the workplace, is also a substantial contributor to PBDE
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exposure (Lorber, 2008).
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Public health concerns have been expressed regarding the potential exposure of humans to
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these compounds (Sjödin et al., 2003; Birnbaum and Staskal, 2004), e.g. effects of pre- and post-
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natal exposure to low PBDE levels on neurodevelopment have been described (Gascon et al., 2011).
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Accordingly, the production and use of penta- and octa-BDE formulations were banned in the
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European Union in 2004. Only deca-BDE is still permitted, despite its use in electronic applications
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was banned in Europe in 2008.
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There is now evidence that BDE-209 also accumulates in humans (Antignac et al., 2009;
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Covaci and Voorspoels, 2005; Fängström et al., 2005; Gómara et al., 2007; Jin et al., 2009; Lunder
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et al., 2010; Sjödin et al., 2008; Takasuga et al., 2004; Thomas et al., 2006; Uemura et al., 2010;
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Vizcaino et al., 2011; Weiss et al., 2006; Zhu et al., 2009), including newborns (Vizcaino et al.,
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2011). However, this compound is rarely measured in human studies from non-exposed
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populations. Limited information is therefore available on its occurrence and effects in human
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populations, although its measurement has been recommended (Hites, 2004).
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Furthermore, many of the available studies on human PBDE exposure are based on
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relatively low samples (n < 250), and often only concerning specific collectives (e.g. mothers), thus
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limiting the possibilities for a comprehensive assessment of the human exposure to these pollutants.
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The NHANES study conducted in the US population in 2003-04 constitutes an exception (n =
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2040), comprising the two genders and a wide age range, although it does not report the
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decabromodiphenyl ether (BDE-209), which is an important limitation. To the best of our
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knowledge, there has not been an equivalent study in Europe to date, serving to complete and
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corroborate it.
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The present study investigates PBDE serum levels in a European general population (that of
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Catalonia, a Mediterranean region in southern Europe) covering both genders, a wide age range and
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diverse socio-demographic conditions, including body mass index (BMI), parity among women,
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birthplace, educational level, and social class. Samples were obtained from a general public health
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survey encompassing 8,400 people, conducted in 2002. Fourteen PBDE congeners including BDE-
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209, were analysed in a subset of 731 individuals. The present study is therefore the second largest
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survey in terms of individual PBDE measurements in a human population, and the first to cover
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specifically BDE-209. The results are then compared with other populations worldwide and
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discussed in terms of socio-demographic determinants.
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2. Materials and methods
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2.1. Population and study design
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The study sample (n = 731) is based on a public health survey (n = 8,400) that was
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conducted by the Government of Catalonia in 2002, including a health exam and blood tests (n =
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2,100). Catalonia is a Mediterranean region of 32,000 km2 in southern Europe, with a total
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population of 6.5 million inhabitants as of 2002. Its major economic activities include agriculture
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and industry.
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The survey provided a valuable and representative sample of the general population in terms
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of age (18-74 years), sex and socio-demographic conditions. Information on body mass index
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(BMI), cholesterol and triglyceride was obtained from the health exam. Information on the
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demographic variables, such as age, sex, place of birth, educational level and parity in women was
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obtained from face-to-face interviews that were conducted between October 2001 and April 2002.
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Social class was estimated through the household occupational status based on the Spanish
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Occupational Classification (Grupo de trabajo SEE-SEMFC, 2000). Further details of the study
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design are available in precedent publications (Juncà et al., 2003; Porta et al., 2010).
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2.2. Laboratory analytical methods
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A total of 14 PBDE congeners were analysed, as follows: BDE-17 and BDE-28
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(tribrominates); BDE-47, BDE-66 and BDE-71 (tetrabrominates); BDE-85, BDE-99 and BDE-100
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(pentabrominates); BDE-153, BDE-154 and BDE-138 (hexabrominates); BDE-183 and BDE-190
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(heptabrominates); and BDE-209 (decabrominates). The analytical methods and quality control
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procedures were standard, as described in detail in a precedent publication (Vizcaino et al., 2009).
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They are summarised next.
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Serum samples (1 ml) were introduced into 10-ml centrifuge tubes. A standard solution
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containing PCB-209 (100 ng/ml; 25 µl) followed by 3 ml of n-hexane and 2 ml of conc. H2SO4
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were added. The suspension was mixed in a vortex (ca. 1,500 rpm, 30 s) and centrifuged (ca. 3,500
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rpm, 5 min). The supernatant n-hexane layer was transferred into a second centrifuge tube using a
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Pasteur pipette. Further n-hexane (2 ml) was added to the first tube containing the H2SO4/serum
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mixture, stirred and then centrifuged. This last step was repeated again yielding a combined extract
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of 7 ml of n-hexane, to which 2 ml of conc. H2SO4 were added; the suspension was then mixed
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(vortex mixer, ca 1500 rpm, 60 s), centrifuged (3500 rpm, 10 min), and the supernatant n-hexane
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was then transferred to a conical bottomed, graduated tube. These n-hexane extracts were reduced
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to near dryness under a stream of pure nitrogen. Then, the solutions were quantitatively transferred
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to GC vials using four 25 μl rinses of isooctane. BDE-118 (20 μl) and [13C]-BDE-209 (10 μl)
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standards were added before injection. The PBDE analysis was performed by gas chromatography
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(GC, Agilent Technologies 6890N) coupled with mass spectrometry (MS, Agilent Technologies
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5975) operating in a negative chemical ionization mode (NICI). The instrument was equipped with
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a low bleed SGE-BPX5 MS fused silica capillary column (length of 15 m, internal diameter of 0.25
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mm, 0.10 μm film thickness).
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One procedural blank was included in each sample batch, with either 9, 15 or 19 samples.
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Method detection limits were calculated at three times the standard deviation of the procedural
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blank levels. Detection limits ranged between 0.0018 and 0.0089 ng/ml, depending on the BDE
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congener. PBDE identification was based on retention time and mass spectral information.
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Quantification was performed by reference to linear calibration curves and correction by the
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standards (Vizcaino et al., 2009). Average recovery of the PCB-209 standard was 64 ±22% (mean ±
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standard deviation). Final validation was made by analysis of proficiency testing materials obtained
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from the Arctic Monitoring and Assessment Program (AMAP Ring Test, 2012). The laboratory
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participates regularly in the AMAP Ring Test Proficiency Program for POPs in human serum and
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the results, including BDE-209 concentrations, usually range within 20% of the consensus values.
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2.3. Data analysis
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Data analysis and graphics were performed using the statistical software R (version 2.15;
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Vienna, 2012). Statistical analyses were focused on the following congeners: 28, 47, 85, 99, 100,
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153, 154 and 209. These compounds were selected because they met detection limits in over 50% of
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samples. For the samples with non-detected and non-quantified values, a score of one-half the LD
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and the LQ was assigned, respectively. Serum concentrations were expressed in a lipid-adjusted
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basis (ng/g lipid) based on levels of cholesterol and triglyceride (Phillips et al., 1989). These levels
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were determined enzymatically, using Txad-Pap and CIN-UV methods, respectively.
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Geometric means (GM) and 95% confidence intervals (CI) have been used for descriptive
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analysis (Figure 1). For the multivariate regression models, concentrations were transformed into
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the natural logarithm in order to normalise the distribution of the values and to avoid violating
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regression assumptions of the normal distribution. The statistical significance of age, sex, place of
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birth, BMI, educational level and social class, as well as parity in women, were assessed for the
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distributions of serum PBDE concentrations using multivariate regression models. For these models
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each of the following variables were categorised into two groups: For age: young (<30 years) and
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adults (>30 years); for place of birth: Spanish-born (including Catalan-born people) and born in
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another country; for parity: women without offspring and women with one or more children; for
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educational level: low education (including individuals without formal education and with primary
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school studies) and high education (comprising studies at secondary school and university); for
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social class: I-IV and V (which is the least affluent one). The BMI was kept in the three original
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groups: normal weight, overweight and obesity. Since all covariates were expressed as factors
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(dummy variables), the exponentials of the coefficients provided the odds ratios (OR) against the
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reference categories. Confidence intervals at 95% were used. Robustness checks were made in
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order to confirm the results of the analyses.
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3. Results and discussion
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3.1. Socio-demographic characteristics of the studied population
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Forty-four percent of the participants were men and fifty-six percent were women (Table 1).
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The mean age was 45 years (standard deviation = 15), ranging between 18 and 74 years. About 70%
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of the total population was born in Catalonia, 26% were of other Spanish origin and only 3% were
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born abroad. The BMI encompassed a large spectrum of cases from underweight (16.9 kg/m2) to
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obesity (52.7 kg/m2): overweight affected 38% of the sample whereas obesity was found in 20%
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(Table 1). Concerning parity, 27% of the women had no descendants, 18% had one child and 55%
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were multiparous. Roughly one third of the studied individuals had a primary school degree, 46%
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had secondary degree and about 12% had a university degree. Almost half of the study population
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was classified within the three more affluent social classes (classes I to III), whereas about 8%
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belonged to the least affluent class (class V) (Table 1). No significant differences in educational or
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social class characteristics were found between the two gender groups.
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3.2. PBDE distributions
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The congener BDE-209 was found above the limit of detection in most cases (83% of the
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cases), followed by congeners BDE-47 and BDE-153 (with 74% and 70%, respectively) (Table 2).
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BDE congeners 28, 85, 99, 100 and 154 were found above limit of detection in around 50-60% of
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the samples. The congeners only found above the limit of detection in less than 30% of the serum
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samples were not included in the calculations of the sections dealing with age, gender, BMI,
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education or social class dependences.
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BDE-209 was the most abundant congener (median 3.7 ng/g lipid; Table 2) accounting for
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28% of total PBDEs. BDE-47 and BDE-99 were the second and third predominant congeners,
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accounting for 18% (median 2.6 ng/g lipid) and 10% (median 1.2 ng/g lipid) of ΣPBDE,
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respectively. Although BDE-99 was found at higher concentration than BDE-153 (1.2 ng/g lipid vs
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0.94 ng/g lipid), it was identified above the limit of detection in less samples (58% versus 70%,
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respectively) (Table 2). The relatively high proportions of BDE-209, BDE-47 and BDE-99 are
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consistent with the predominance of the commercial mixtures deca-BDE and penta-BDE in the
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studied samples. However, in the original composition of penta-BDE, the congener BDE-99
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predominates over BDE-47 (Alaee et al., 2003).
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The dominance of the BDE-209 congener in PBDE distributions from human populations
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has also been observed in a study on French women (Antignac et al., 2009), in two studies of
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Japanese populations (Takasuga et al., 2004; Uemura et al., 2010) and in two studies of Chinese
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populations (Jin et al., 2009; Zhu et al., 2009). BDE-47 has been observed to predominate the
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congener distributions of most populations in which BDE-209 was not measured: e.g. studies in
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North America (Castorina et al., 2011; Johnson et al., 2010; Sandanger et al., 2007; Sjödin et al.,
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2001, 2008), New Zealand (Harrad and Porter, 2007), Asia (Bi et al., 2006; Lee et al., 2007) and
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Europe, including studies conducted in Sweden, Norway, Germany and the Faroe Islands
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(Fängström et al., 2005; Guvenius et al., 2003; Schröter-Kermani et al., 2000; Thomsen et al.,
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2002). In some cases in which BDE-209 was included, BDE-47 was also found to predominate over
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the decabromate congener (Gómara et al., 2007; Vizcaino et al., 2011). In a few European studies
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performed in British (Thomas et al., 2006), Belgian (Roosens et al., 2010) and Greek (Kalantzi et
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al., 2011) populations, BDE-153 was found to be the dominant PBDE congener. The second of
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these studies did not include BDE-209. These differences in congener profile may reflect exposure
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to different commercial PBDE formulations, such as the penta, octa and deca-BDE mixtures
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mentioned above.
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Moreover, environmental processes (Bezares-Cruz et al., 2004; Lacorte et al., 2003) and
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metabolic transformations (Bartrons et al., 2011, 2012; Robrock et al., 2008) may also influence on
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the final mixtures of accumulated PBDEs. In this respect, the studied population of Catalonia
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showed showed a relatively high abundance of BDE-47 and BDE-28 (Table 2), compared to the
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proportion of BDE-99 in the penta-BDE commercial mixtures (Alaee et al., 2003). Some metabolic
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or environmental effects (Alaee et al., 2003; Bartrons et al., 2012; Robrock et al., 2008) may be
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responsible for these enhanced concentrations of the tri- and tetrabromo congeners. Similarly, the
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high relative proportions of BDE-153 and BDE-154 in the studied population (Table 2), compared
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to BDE-183, which is the dominant congener in the octa-BDE commercial mixtures (Alaee et al.,
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2003), may also reflect debromination of more brominated BDE congeners and the bioaccumulation
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of these hexabromo congeners. In fact, metabolic enrichment of BDE-153 upon exposure to another
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PBDE, namely BDE-209, has been observed in mammals (e.g. polar bears; Sormo et al., 2006) and
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fish (Kierkegaard et al., 1999), involving increases of both BDE-153 and BDE-154 in the latter.
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3.3. PBDE concentrations
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ΣPBDE, encompassing the 14 congeners analysed, ranged between 2.2 ng/g lipid and 150
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ng/g lipid (Table 2), with a median value of 15 ng/g lipid (Table 3). BDE-209 concentrations ranged
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between 0.45–50 ng/g lipid, with median and mean values of 3.7 ng/g lipid and 4.6 ng/g lipid,
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respectively (Tables 2 and 3). This congener shows the highest concentration (Table 3), which
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suggests that the deca-BDE commercial mixture is a major source of exposure in the population of
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the study.
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Compared to other European studies, the analysed Catalan population has the highest
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ΣPBDE concentrations (medians ranging between 2.1 ng/g lipid in Sweden and 5.6 ng/g lipid in the
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UK), even above other populations studied in Spain (medians 11 and 9.6 ng/g lipid) (Table 3). Part
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of the difference is due to the high proportion of BDE-209 in the Catalan population. Only two
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European studies have shown a dominance of BDE-209 in the total BDE mixture: France (median
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5.8 ng/g lipid; Antignac et al., 2009) and Belgium (median 11.1 ng/g lipid; Covaci and Voorspoels,
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2005) (Table 3); yet in these two cases, particularly in the second, the number of individuals
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examined is comparatively small and thus any comparison should be done with caution.
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The commercial deca-BDE mixture is still allowed for use in Europe, for applications other
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than electrical and electronic equipment. In fact, BDE-209 remains a major environmental pollutant
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in Europe. It has been found to largely dominate the PBDE congener distributions in remote sites
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such as high mountain lakes in the Pyrenees and the Tatras (Bartrons et al., 2011). This compound is
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also the dominant congener in coastal effluents into the Mediterranean coast (Eljarrat el al., 2005;
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Salvado et al., 2012), which reflects a strong influence from the use of the deca-BDE mixture in the
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basins of the studied river mouths. BDE-209 is also very abundant, sometimes the highest, in the
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PBDE congener distributions from European commercial foodstuffs (Domingo et al., 2006; Gómara
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et al., 2006).
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Because of the scarcity of BDE-209 measurements, the comparison of the PBDE
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concentrations between populations sometimes requires the use of other ΣPBDE values such as the
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sum of tri- to hepta-BDE congeners (ΣPBDE3_7) (Tables 2 and 3). BDE-47 is also used (Table 3)
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due to the abundance of this congener in human samples. The median ΣPBDE3_7 in the Catalan
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population is 8.0 ng/g lipid (Table 3). This value is still higher than all previous studies reported on
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European populations, except one in Spain, which has a median of 8.5 ng/g lipid (Gómara et al.,
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2007). Nevertheless, the median BDE-47 in the Catalan population study is at 2.6 ng/g lipid, the
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second highest after the reported value in the French population (2.8 ng/g lipid) (Table 3). However,
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other European studies have found similar BDE-47 median values; e.g. in Spain (2.3 and 2.5 ng/g
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lipid) and in one study in Norway (2.2 ng/g lipid) (Table 3).
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The comparison of the median values of other congeners within a European context show in
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the Catalan population high concentrations of BDE-85, BDE-99 and BDE-100, which is consistent
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with a higher exposure to the commercial penta-BDE mixture. The sample collection for this study
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was performed in 2002, whereas the manufacture and use of this PBDE mixture was only banned in
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Europe two years later, in 2004. These dates are therefore consistent with significant exposure of
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the study population to this commercial product.
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The ΣPBDE concentrations in Australia are similar to those found in Catalonia, with 16 ng/g
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lipid (mean) and 15.4 ng/g lipid (median), respectively. In contrast, the New Zealand values are
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lower: the ΣPBDE3_7 was 6.1 ng/ lipid versus 8.0 ng/g lipid in Catalonia (Table 3). However,
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comparison of the BDE-47 concentrations shows lower values in the Catalan population (2.6 ng/g
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lipid) than in those from Australia (8.4 ng/g lipid) or New Zealand (3.2 ng/g lipid).
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The comparison of the results from the present study with those reported from Asian
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populations show similar median ΣPBDE concentrations in Korea, but higher levels in Catalonia
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than in China and Japan (Table 3). However, it is likely that the real PBDE concentrations in the
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Korean population are actually higher than in the Catalan study since the BDE-209 concentrations
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were not assessed. The BDE-47 median concentrations were remarkably different between the
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Catalan and the Korean populations (2.6 ng/g lipid versus 5.2 ng/g lipid, respectively). One of the
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Japanese studies shows a higher median concentrations of BDE-209 than in the Catalan population
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(6.9 ng/g lipid versus 3.7 ng/g lipid, respectively) (Takasuga et al., 2004). In fact, the PBDE
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distribution in this Japanese population study is strongly dominated by this congener (73%) which
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is uncommon since the proportion of BDE-209 in ΣPBDE from general populations studied in other
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countries usually range from 0 and 25% (Table 3). The high values reported in one Chinese study
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(mean ΣPBDE: 613 ng/g lipid; BDE-209: 403 ng/g lipid; BDE-47: 21 ng/g lipid) (Jin et al., 2009)
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correspond to a resident population in a PBDE-production area.
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All population studies from the USA, except one (Turyk et al., 2010), reported higher
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median ΣPBDE (17-39 ng/g lipid) than the Catalan study (15 ng/g lipid) (Table 3). Data in one of
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these US studies has been obtained from analysis of a very large series of individuals (n= 2040;
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Sjödin et al., 2008). In Canada, the reported median ΣPBDE3_7 is also higher than in Catalonia (13.4
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versus 8.0 ng/g lipid) (Table 3). Differences between the populations of the USA and Canada, on
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the one hand, and the population of Catalonia, on the other, are also observed when comparing
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BDE-47 concentrations: median of 8.8-34 ng/g lipid in the USA (excluding the value reported by
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Turyk et al. (2010)) and geometric mean of 8.1 ng/g lipid in Canada versus a much lower median
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of 2.6 ng/g lipid in Catalonia (Table 3). The causes for the higher PBDE concentrations in US
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populations, compared to European levels, have been already discussed (Hites, 2004).
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3.4. Sex differences
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The observed ΣPBDE concentrations were similar between women and men. The latter
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tended to have slightly higher levels, but the differences were not statistically significant (Figure 1).
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Studies diverge: some have found higher PBDE levels in men (Covaci et al., 2008; Schöter-
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Kermani et al., 2000; Sjödin et al., 2008; Takasuga et al., 2004; Thomsen et al., 2002; Uemura et al.,
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2010; Zota et al., 2008), others in women (Jin et al., 2009; Schecter et al., 2005) and others have not
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found any difference (Harrad and Porter, 2007; Anderson et al., 2008). Higher concentrations of
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PBDEs in men than in women may result from differences in diet, occupational exposures or
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metabolism (Robrock et al., 2008). In fact, differences in exposure sources, whether at home or at
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the work place, may play a relevant role (Covaci et al., 2008). Higher BDE-209 levels in women
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have been related to dermal exposure in computer-rich environments, such as office desks
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(Takasuga et al., 2004). In contrast, the lower concentrations of other PBDEs in women could be
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explained by placental transfer and by detoxification during lactation (Takasuga et al., 2004; Toms
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et al., 2009). Some studies have found statistically significant sex differences for BDE-153 (Sjödin
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et al., 2003; Toms et al., 2009). Furthermore, serum concentrations of BDE-153 were not correlated
359
between males and females in a study specifically considering the PBDE concentrations within
360
couples (Johnson et al., 2010). In the present study, the multivariate regression did not show
361
statistically significant trends for gender.
362
363
3.5. Age dependence
364
In the studied population, the most notable observation was that young people (<30 years
365
old) showed higher GM concentrations of most PBDE congeners than adults (>30 years old). The
366
differences were particularly relevant for the most abundant congeners, namely BDE-47, BDE-99
367
and BDE-209 (Figure 1), with differences of 46%, 31% and 14%, respectively. These concentration
368
differences were maintained in the multivariate regression models after adjusting for other
369
covariates (Figure 2). In fact, age was clearly the most significant variable for the description of the
370
distribution of PBDE concentrations in the study population.
371
However, most previous reports did not observe changes in PBDE concentrations in relation
372
to age (Antignac et al., 2009; Harrad and Porter, 2007; Jin et al., 2009; Lee et al., 2007; Sjödin et
373
al., 2000; Thomsen et al., 2002; Zota et al., 2008). These results are in contrast with those from
374
other compounds having similar physical-chemical properties such as polychlorobiphenyls or
375
polychlorinated pesticides, which tend to increase with age (Vizcaino et al., 2010). In the case of
376
PBDEs, their relatively recent manufacture and use, since around the 1980s, implies shorter times
377
of population exposure and hence are often cited to explain the absence of age-dependence
378
observations. However, methodological aspects may also influence such observations since many
379
studies have low sample sizes or were performed on specific population groups over short age
380
intervals (e.g. mothers during the fertile period).
381
In fact, the two studies involving large population databases also found higher
382
concentrations of some PBDE congeners among the youngest individuals. The study NHANES in
383
the US (2,040 PBDE determinations) shows the highest concentrations of BDE-47, BDE-99, BDE-
384
100 and BDE-153 in the age interval of 12-19 years, followed by the interval of 20-29 years;
385
whereas older age periods showed approximately constant concentrations (Sjödin et al., 2008). An
386
Australian study (80 PBDE determinations from age-interval pooled samples taken from a studied
387
population of 8,132 individuals) showed higher concentrations of BDE-47, BDE-99, BDE-100 and
388
BDE-153 in infants (0-4 years) and children (5-15 years) than in adults (Toms et al., 2008, 2009).
389
The results from the population in Catalonia concur with studies involving large databases.
390
Based on a wide age interval (18-74 years), the study shows the highest concentrations in the
391
youngest age interval (<30 years) and, when shorter time periods were assessed, higher
392
concentrations are found in the 18-23 year interval than in the 24-29 years (Figure 3). The analysis
393
of BDE-209 provides additional evidence of this unusual trend, showing how a highly brominated
394
highly insoluble congener has also higher concentrations in the age group under 30 years.
395
Whereas the accumulation of chlorinated pollutants usually increase with age (Vizcaino et
396
al., 2010), data suggest that PBDE accumulation has an opposite age-dependence trend. Since both
397
compound groups have similar physical-chemical properties, one would predict a similar pattern of
398
accumulation. The reason may lie in their different human history, as PBDEs have been
399
manufactured and polluting human populations since most recent times. The fact that BDE-209, the
400
compound with most dissimilar physical-chemical properties, also shows the same trend as the
401
other less-brominated BDE congeners, reinforces this conclusion. Lifestyle and activity differences
402
between young and old individuals have been quoted to also explain the higher accumulation of
403
PBDEs in the former (Sjödin et al., 2008). In addition, other aspects so far undemonstrated may also
404
be contributing, such as pharmacokinetic differences in young and adult people or the effects of
405
growth in environments with high PBDE concentrations. In fact, children are more prone to
406
incidental dust ingestion, and it is worth mentioning that precisely the youngest individuals of the
407
present study – who were born between 1972 and 1984 – were newborns and children when PBDEs
408
began to be manufactured.
409
In the present study, the observed threshold of 30 years and the time period of sample
410
collection (i.e. year 2002) implies that adults (who show lower concentrations ) were born in the
411
1970s or earlier, a time period of low use of material treated with PBDEs. Meanwhile, the group of
412
young participants were born and grew in the 1980s or later, which is precisely the time period of
413
scaling up the use of materials containing PBDEs. The present study (from a European population),
414
as well as the studies in the US (Sjödin et al., 2008) and Australia (Toms et al., 2009), are based on
415
samples collected in about the same time periods. Therefore the common trend of an inverse age-
416
dependence of PBDE accumulation (higher under the 30 year threshold) is correlated to the period
417
of rapid use of such compounds and thus suggest increasing population exposure to these
418
compounds (i.e. the overall PBDE concentrations in the general population will also further
419
increase). Accordingly, public health and environmental agencies should encourage research on
420
these compounds in anticipation of the effects of these increasing concentrations.
421
422
3.6. Body mass index
423
Individuals with normal weight showed slightly higher concentrations of BDE-47 and BDE-
424
209 than individuals with overweight or obesity (Figure 1). Lower concentrations of BDE-153 in
425
obese people have been also observed in other studies (Weiss et al., 2006). However, previous
426
literature information does not show a clear trend, even in some cases no association between BMI
427
and PBDE has been reported (Sandanger et al., 2007), and there are also reports of positive
428
associations (Anderson et al., 2008; Castorina et al., 2011). The observed lower concentrations of
429
BDE-47 and BDE-209 in the normal weight individuals of the study population are consistent with
430
lower past exposures to these compounds. Thus, weight increases usually occur over long time
431
periods and lower past environmental or domestic PBDE concentrations likely involved lower
432
PBDE/weight rates in the early stages of weight increase. Nevertheless, the multivariate regression
433
models of the population considered in the present study showed that the obesity group scored
434
lower in BDE-47, BDE-99 and BDE-209 but that the results were not statistically significant (p <
435
0.05) (Figure 2).
436
437
3.7. Place of birth
438
Unadjusted analysis showed that the individuals who were born abroad (outside Catalonia
439
and Spain) had higher GM levels of BDE-47, BDE-99 and BDE-100. Conversely, they had lower
440
concentrations of BDE-209 (Figure 1). In this respect, the GM concentrations of this last congener
441
were slightly higher among Catalan-born individuals than in Spanish-born outside Catalonia or
442
abroad. These differences could reflect specific PBDE exposures in different geographic areas, e.g.
443
higher contamination by deca-BDE in the individuals born in Catalonia and higher contamination
444
by penta-BDE in the individuals of the population born abroad Spain. However, these differences
445
were not statistically significant since application of the multivariate regression models eliminated
446
place of birth as descriptive variable for the PBDE differences.
447
448
3.8. Parity
449
Women without offspring showed higher GM BDE-47 concentrations than women with
450
children. The difference was not so well defined for the other PBDE congeners or between women
451
having 1, 2 or ≥3 children. Lower concentrations of organochlorine compounds have been found in
452
multiparous compared to primiparous women, since delivery and breastfeeding provides a way of
453
eliminating part of the burden of these pollutants (Polder et al., 2009). Thus, lower PBDE
454
concentrations would be expected among women with children according to the similar physical-
455
chemical properties of these compounds. Nevertheless the multivariate models show that the
456
influence of parity in the total burden of BDE-47 and the other BDE congeners is not statistically
457
significant among women in the studied population (Figure 2).
458
459
3.9. Educational level
460
Educational level did not result into statistically significant differences in PBDE
461
accumulation in the population of study (Figures 1 and 2). Some studies have reported positive
462
associations between individuals with higher education and higher PBDE concentrations (Castorina
463
et al., 2011) while others have found that less educated individuals had higher PBDE levels (Zota et
464
al., 2008). The matter remains inconclusive.
465
466
3.10. Social class
467
Individuals belonging to the less affluent social class (group V) showed higher
468
concentrations of BDE-47, compared to the rest of the population (Figure 1). Yet this difference was
469
only observed for the least affluent social class, being retained in the multivariate model as
470
statistically significant (Figure 2). However, it was not significant at all for the other major
471
congeners, e.g. BDE-99 and BDE-209 (Figures 1 and 2). These results were in agreement with
472
some previous observations. A statistically significant difference (p < 0.01) between ΣPBDE
473
concentrations (BDE-209 not included) and household income was identified among the US
474
population examined in the NHANES study (Zota et al., 2008), involving higher concentrations for
475
income ≤US$20,000. Similarly, in another study on US consumers of sport-caught fish (Anderson
476
et al., 2008), individuals with annual income <$US45,000 were observed to have higher ΣPBDE
477
serum concentrations (BDE-209 not included). Differences in lifestyle, quality of the domestic
478
furniture and other home items and mobility may be responsible for these higher concentrations
479
among lower social class individuals.
480
481
482
4. Conclusions
483
484
In summary, age appears to be the most important socio-demographic characteristic to
485
explain the distribution of several PBDEs in the population studied. In essence, young individuals
486
showed the highest concentrations, which seems related to high exposure to these pollutants during
487
their growth period, which is a period of higher metabolism than adults, coinciding with the period,
488
around the 1980s, when most PBDEs started to be manufactured and traded. This effect actually
489
overcomes the bioaccumulation trend through time, consisting in the higher accumulation in older
490
ages due to the lipophilic character and chemical stability of these compounds (such as it is often
491
observed in studies on human accumulation of polychlorinated pollutants). The socio-economic
492
level is also a factor of PBDE concentrations in the present studied population, but only observed to
493
be relevant for the least affluent class and for some congeners and may be related to domestic
494
environmental aspects.
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497
498
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500
501
Acknowledgements
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This study was funded by the Department of Health of the Government of Catalonia, the
504
Foundation La Marató de TV3 (grant No. 090431), the Spanish Ministry of Science and Innovation
505
(Consolider Ingenio GRACCIE, CSD2007-00067) and the EU funded project ArcRisk (FP7-ENV-
506
2008-1-226534). The paper was also sponsored by the research group 2009SGR1178 from the
507
Government of Catalonia. M.C. is grateful for a grant from the JAE Predoctoral program (CSIC). R.
508
Chaler, D. Fanjul and M. Comesaña are thanked for their kind support with the GC-MS analyses.
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References
Alaee, M.; Arias, P.; Sjodin, A.; Bergman, A. 2003 An overview of commercially used brominated
flame retardants, their applications, their use patterns in different countries/regions and
possible modes of release. Environ Int 29, 683-689.
Anderson, H. A.; Imm, P.; Knobeloch, L.; Turyk, M.; Mathew, J.; et al. 2008 Polybrominated
diphenyl ethers (PBDE) in serum: findings from a US cohort of consumers of sport-caught
fish. Chemosphere 73, 187–194.
Antignac, J.-P.; Cariou, R.; Zalko, D.; Berrebi, A.; Cravedi, J.-P.; et al. 2009 Exposure assessment of
French women and their newborn to brominated flame retardants: determination of tri- to
deca-polybromodiphenylethers (PBDE) in maternal adipose tissue, serum, breast milk and
cord serum. Environ Pollut 57, 164–173.
Bartrons, M.; Grimalt, J. O.; Catalan, J. 2011 Altitudinal distributions of BDE-209 and other
polybromodiphenyl ethers in high mountain lakes. Environ Pollut 159, 1816-1822.
Bartrons, M.; Grimalt, J. O.; de Mendoza, G.; Catalan, J. 2012 Pollutant dehalogenation capability
may depend on the trophic evolutionary history of the organism: PBDEs in freshwater food
webs. Plos One, 7, e41829
Bezares-Cruz, J.; Jafvert, C. T.; Hua, I. 2004 Solar photodecomposition of decabromodiphenyl
ether: products and quantum yield. Environ Sci Technol. 38, 4149-4156.
531
532
533
534
535
536
537
538
539
540
541
542
543
544
545
546
547
548
549
550
551
552
553
554
555
556
557
558
559
560
561
562
563
564
565
566
567
568
569
570
571
572
573
574
575
576
577
578
579
580
581
582
Bi, X.; Qu, W.; Sheng, G.; Zhang, W.; Mai, B.; et al. 2006 Polybrominated diphenyl ethers in South
China maternal and fetal blood and breast milk. Environ Pollut, 144, 1024–1030.
Birnbaum, L. S.; Staskal, D. F. 2004 Brominated flame retardants: cause for concern? Environ
Health Perspect, 112, 9–17.
Bocio, A.; Llobet, J. M.; Domingo, J. L.; Corbella, J.; Teixidó, A.; Casas, C. 2003 Polybrominated
diphenyl ethers (PBDEs) in foodstuffs: human exposure through the diet. J Agric Food
Chem, 3191–3195.
Castorina, R.; Bradman, A.; Sjödin, A.; Fenster, L.; Jones, R. S.; et al. 2011 Determinants of serum
polybrominated diphenyl ether (PBDE) levels among pregnant women in the CHAMACOS
cohort. Environ Sci Technol, 45, 6553–6560.
Centre de Toxicologie. Institut National de Santé Publique du Québec, 2012 AMAP Ring Test
Proficiency Program; http://www.inspq.qc.ca/ctq/paqe/amap/Default.asp?Page=1&Lg=en
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 Analyt Technol Biomed Life
Sci, 827, 216–223.
Covaci, A.; Voorspoels, S.; Roosens, L.; Jacobs, W.; Blust, R.; Neels, H. 2008 Polybrominated
diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in human liver and adipose
tissue samples from Belgium. Chemosphere, 73, 170-175.
Domingo, J. L.; Bocio, A.; Falcó, G.; Llobett, J. M. 2006 Exposure to PBDEs and PCDEs
associated with the consumption of edible marine species. Environ Sci Technol, 40, 4394–
4399.
Eljarrat, E.; Cal, A. D. L.; Larrazabal, D.; Fabrellas, B.; Fernandez-Alba, A. R.; et al. 2005
Occurrence of polybrominated diphenylethers, polychlorinated dibenzo-p-dioxins,
dibenzofurans and biphenyls in coastal sediments from Spain. Environ Pollut, 136, 493–501.
Fängström, B.; Hovander, L.; Bignert, A.; Athanassiadis, I.; Linderholm, L.; et al. 2005
Concentrations of polybrominated diphenyl ethers, polychlonnated biphenyls, and
polychlorobiphenylols in serum from pregnant Faroese women and their children 7 years
later. Environ Sci Technol, 39, 9457–9463.
Fraser, A. J.;Webster, T. F.; McClean, M. D. 2009 Diet contributes significantly to the body burden
of PBDEs in the general U.S. population. Environ Health Perspect, 117, 1520–1525.
Gascon, M.; Vrijheid, M.; Martínez, D.; Forns, J.; Grimalt, J.O.; et al. 2011 Effects of pre and
postnatal exposure to low levels of polybromodiphenyl ethers on neurodevelopment and
thyroid hormone levels at 4 years of age Environ Int, 37, 605-611.
Gómara, B.; Herrero, L.; González, M. J. 2006 Survey of polybrominated diphenyl ether levels in
Spanish commercial foodstuffs. Environ Sci Technol, 40, 7541–7547.
Gómara, B.; Herrero, L.; Ramos, J. J.; Mateo, J. R.; Fernández, M. A.; et al. 2007 Distribution of
polybrominated diphenyl ethers in human umbilical cord serum, paternal serum, maternal
serum, placentas, and breast milk from Madrid population, Spain. Environ Sci Technol, 41,
6961–6968.
Grupo de trabajo de la Sociedad Española de Epidemiología y de la Sociedad Española de Medicina
Familiar y Comunitaria. 2000 Una propuesta de medida de la clase social. Atención
Primaria, 25, 350–363.
Guardia, M. J. L.; Hale, R. C.; Harvey, E. 2006 Detailed Polybrominated Diphenyl Ether (PBDE)
Congener Composition of the Widely Used Penta-, Octa-, and Deca-PBDE Technical
Flame retardant Mixtures. Environmental Science and Technology, 40, 6247–6254.
Guvenius, D. M.; Aronsson, A.; Ekman-Ordeberg, G.; Bergman, A.; Norén, K. 2003 Human
prenatal and postnatal exposure to polybrominated diphenyl ethers, polychlorinated
biphenyls, polychlorobiphenylols, and pentachlorophenol. Environ Health Perspect, 111,
1235–1241.
Harrad, S.; Porter, L. 2007 Concentrations of polybrominated diphenyl ethers in blood serum from
New Zealand. Chemosphere, 66, 2019–2023.
583
584
585
586
587
588
589
590
591
592
593
594
595
596
597
598
599
600
601
602
603
604
605
606
607
608
609
610
611
612
613
614
615
616
617
618
619
620
621
622
623
624
625
626
627
628
629
630
631
632
633
634
Hites, R. A. 2004 Polybrominated diphenyl ethers in the environment and in people: a metaanalysis
of concentrations. Environ Sci Technol, 38, 945–956.
Jin, J.; Wang, Y.; Yang, C.; Hu, J.; Liu, W.; et al. 2009 Polybrominated diphenyl ethers in the serum
and breast milk of the resident population from production area, China. Environ Int, 35,
1048–1052.
Johnson, P. I.; Stapleton, H. M.; Sjodin, A.; Meeker, J. D. 2010 Relationships between
polybrominated diphenyl ether concentrations in house dust and serum. Environ Sci
Technol, 44, 5627–5632.
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.
Juncà, S.; Guillén, M.; Aragay, J. M.; Brugulat, P.; Castell, C.; et al. 2003 Methodological aspects in
the evaluation of health and risk-reduction objectives of Health Plan for Catalonia for the
year 2000. Med Clin (Barc), 121 Suppl 1, 10–19.
Kalantzi, O. I.; Geens, T.; Covaci, A.; Siskos, P. A. 2011 Distribution of polybrominated diphenyl
ethers (PBDEs) and other persistent organic pollutants in human serum from Greece.
Environ Int, 37, 349-353.
Kierkegaard, A.; Balk, L.; Sellstrom, V.; Tjamlund, U.; Om, U. et al. 1999 Dietary uptake and
biological effects of decabromodiphenyl ether in rainbow trout (Oncorhynchus myskiss).
Env Sci Technol, 33, 1612-1617.
Lacorte, S.; Guillamon, M.; Martinez, E.; Viana, P.; Barceló, D. 2003 Occurrence and specific
congener profile of 40 polybrominated diphenyl ethers in river and coastal sediments from
Portugal. Environ Sci Technol, 37, 892-898.
Lee, S.-J.; Ikonomou, M. G.; Park, H.; Baek, S.-Y.; Chang, Y.-S. 2007 Polybrominated diphenyl
ethers in blood from Korean incinerator workers and general population. Chemosphere, 67,
489–497.
Lorber, M. 2008 Exposure of Americans to polybrominated diphenyl ethers. J Expo Sci Environ
Epidemiol, 18, 2-19.
Lunder, S.; Hovander, L.; Athanassiadis, I.; Bergman, A. 2010 Significantly higher polybrominated
diphenyl ether levels in young U.S. children than in their mothers. Environ Sci Technol, 44,
5256–5262.
Phillips, D. L.; Pirkle, J. L.; Burse, V. W.; Bernert, J. T.; Henderson, L. O.; Needham, L. L. 1989
Chlorinated hydrocarbon levels in human serum: effects of fasting and feeding. Arch
Environ Contam Toxicol, 18, 495–500.
Polder, A.; Skaare, J. U.; Skjerve, E.; Løken, M.; Eggesbø, M. 2009 Levels of chlorinated pesticides
and polychlorinated biphenyls in Norwegian breast milk (2002-2006), and factors that may
predict the level of contamination. Sci Total Environ, 407:4584-4590
Porta, M.; Gasull, M.; Puigdomènech, E.; Garí, M.; de Basea, M. B.; et al. 2010 Distribution of
blood concentrations of persistent organic pollutants in a representative sample of the
population of Catalonia. Environ Int, 36, 655–664.
R Development Core Team. 2012 R: A Language and Environment for Statistical Computing. R
Foundation for Statistical Computing: Vienna, Austria; ISBN 3-900051-07-0 available at
http://www.R-project.org/.
Robrock, K. R., Korytar, P., Alvarez-Cohen, L. 2008 Pathways for the anaerobic microbial
debromination of polybrominated diphenyl ethers. Environ Sci Technol, 42, 2845-2852.
Roosens, L.; D’Hollander, W.; Bervoets, L.; Reynders, H.; Campenhout, K. V.; et al. 2010
Brominated flame retardants and perfluorinated chemicals, two groups of persistent
contaminants in Belgian human blood and milk. Environ Pollut, 158, 2546–2552.
Salvado, J. A.; Grimalt, J. O.; Lopez, J. F.; Durrieu de Madron, X.; Heussner, S.; Canals, M. 2012
Transformation of PBDE mixtures during sediment transport and resuspension in marine
environments (Gulf of Lion, NW Mediterranean Sea). Env Pollut, 168, 87-95.
Sandanger, T. M.; Sinotte, M.; Dumas, P.; Marchand, M.; Sandau, C. D.; et al. 2007 Plasma
635
636
637
638
639
640
641
642
643
644
645
646
647
648
649
650
651
652
653
654
655
656
657
658
659
660
661
662
663
664
665
666
667
668
669
670
671
672
673
674
675
676
677
678
679
680
681
682
683
684
685
686
concentrations of selected organobromine compounds and polychlorinated biphenyls in
postmenopausal women of Québec, Canada. Environ Health Perspect, 115, 1429–1434.
Schecter, A.; Päpke, O.; Tung, K. C.; Joseph, J.; Harris, T. R.; Dahlgren, J. 2005 Polybrominated
diphenyl ether flame retardants in the U.S. population: current levels, temporal trends, and
comparison with dioxins, dibenzofurans, and polychlorinated biphenyls. J Occup Environ
Med, 47, 199–211.
Schröter-Kermani, C.; Helm, D.; Herrmann, T.; Päpke, O. 2000 The German environmental
specimen bank - application in trend monitoring of polybrominated diphenyl ethers in
human blood. Organohalogen Compounds, 47, 49–52.
Sjödin, A.; Hagmar, L.; Klasson-Wehler, E.; Björk, J.; Bergman, A. 2000 Influence of the
consumption of fatty Baltic Sea fish on plasma levels of halogenated environmental
contaminants in Latvian and Swedish men. Environ Health Perspect, 108, 1035–1041.
Sjödin, A.; Patterson, D. G.; Bergman, A. 2001 Brominated flame retardants in serum from U.S.
blood donors. Environ Sci Technol, 35, 3830–3833.
Sjödin, A.; Patterson, D. G.; Bergman, A. 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.; et al. 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.
Sjödin, A.; Wong, L.-Y.; Jones, R. S.; Park, A.; Zhang, Y.; et al. 2008 Serum concentrations of
polybrominated diphenyl ethers (PBDEs) and polybrominated biphenyl (PBB) in the United
States population: 2003-2004. Environ Sci Technol, 42, 1377–1384.
Sormo, E. G.; Salmer, M. P.; Jenssen, B. M.; Hop, H.; Baek, K. et al. 2006 Biomagnification of
polybrominated diphenyl ether and hexabromocyclododecane flame retardants in the polar
bear food chain in Svalbard, Norway. Env Toxicol Chem, 25, 2502-2511.
Takasuga, T.; Senthilkumar, K.; Takemori, H.; Ohi, E.; Tsuji, H.; Nagayama, J. 2004 Impact of
fermented brown rice with Aspergillus oryzae (FEBRA) intake and concentrations of
polybrominated diphenylethers (PBDEs) in blood of humans from Japan. Chemosphere, 57,
795–811.
Thomas, G. O.; Wilkinson, M.; Hodson, S.; Jones, K. C. 2006 Organohalogen chemicals in human
blood from the United Kingdom. Environ Pollut, 141, 30–41.
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.
Toms, L.-M. L.; Harden, F.; Paepke, O.; Hobson, P.; Ryan, J. J.; Mueller, J. F. 2008 Higher
accumulation of polybrominated diphenyl ethers in infants than in adults. Environ Sci
Technol, 42, 7510–7515.
Toms, L.-M. L.; Sjödin, A.; Harden, F.; Hobson, P.; Jones, R.; et al. 2009 Serum polybrominated
diphenyl ether (PBDE) levels are higher in children (2-5 years of age) than in infants and
adults. Environ Health Perspect, 117, 1461–1465.
Turyk, M. E.; Anderson, H. A.; Steenport, D; Buelow, C; Imm, P; Knobeloch, L. 2010 Longitudinal
biomonitoring for polybrominated diphenyl ethers (PBDEs) in residents of the Great Lakes
basin. Chemosphere, 81, 517-522.
Uemura, H.; Arisawa, K.; Hiyoshi, M.; Dakeshita, S.; Kitayama, A.; et al. 2010 Congener-specific
body burden levels and possible determinants of polybrominated diphenyl ethers in the
general Japanese population. Chemosphere, 79, 706–712.
Vizcaino, E.; Arellano, L.; Fernandez, P.; Grimalt, J. O. 2009 Analysis of whole congener mixtures
of polybromodiphenyl ethers by gas chromatography-mass spectrometry in both
environmental and biological samples at femtogram levels. J Chromatogr A, 1216, 5045–
5051.
Vizcaino, E.; Grimalt, J. O.; Lopez-Espinosa, M.-J.; Llop, S.; Rebagliato, M.; Ballester, F. 2010
Maternal origin and other determinants of cord serum organochlorine compound
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688
689
690
691
692
693
694
695
696
697
698
699
700
701
702
703
704
705
concentrations in infants from the general population. Environ Sci Technol, 44, 6488-6495.
Vizcaino, E.; Grimalt, J. O.; Lopez-Espinosa, M.-J.; Llop, S.; Rebagliato, M.; Ballester, F. 2011
Polybromodiphenyl ethers in mothers and their newborns from a non-occupationally
exposed population (Valencia, Spain). Environ Int, 37, 152–157.
Weiss, J.; Wallin, E.; Axmon, A.; Jönsson, B. A. G.; Akesson, H.; et al. 2006 Hydroxy-PCBs,
PBDEs, and HBCDDs in Serum from an Elderly Population of Swedish Fishermen’s Wives
and Associations with Bone Density. Environ Sci Technol, 40, 6282–
6289.
Zhu, L.; Ma, B.; Hites, R. A. 2009 Brominated flame retardants in serum from the general
population in northern China. Environ Sci Technol, 43, 6963–6968.
Zota, A. R.; Rudel, R. A.; Morello-Frosch, R. A.; Brody, J. G. 2008 Elevated house dust and serum
concentrations of PBDEs in California: unintended consequences of furniture flammability
standards? Environ Sci Technol, 42, 8158–8164.
706
707
708
Table 1
Socio-demographic characteristics of the population studied.
All
Men
Women
n (%)
n (%)
n (%)
731 (100) 324 (44)
407 (56)
18-29
138 (19)
55 (17)
83 (21)
30-44
220 (30)
97 (30)
123 (30)
45-59
228 (31)
101 (31)
127 (31)
60-74
145 (20)
71 (22)
74 (18)
Catalonia
518 (71)
235 (73)
283 (70)
Other Spanish areas
189 (26)
79 (24)
110 (27)
24 (3)
10 (3)
14 (3)
Normal weight
314 (43)
108 (33)
206 (50)
Overweight
276 (38)
155 (48)
121 (30)
Obesity
141 (19)
61 (19)
80 (20)
All participants
Age (years)
Place of birth
Other countries
Body mass index
Parity (women)
0
110 (27)
1
72 (18)
2
145 (35)
≥3
80 (20)
Educational level
WFEa
111 (15)
44 (14)
67 (16)
Primary school
198 (27)
82 (25)
116 (29)
Secondary school (I)
177 (24)
83 (26)
94 (23)
Secondary school (II)
158 (22)
76 (23)
82 (20)
University
87 (12)
39 (12)
48 (12)
I (most affluent)
73 (10)
31 (10)
42 (10)
II
85 (12)
36 (11)
49 (12)
III
184 (25)
95 (29)
89 (22)
IV
328 (45)
137 (42)
191 (47)
61 (8)
25 (8)
36 (9)
Social class
V (least affluent)
709
710
a
Without formal education.
711
712
713
714
Table 2
Serum PBDE concentrations (ng/g lipid) of the BDE congeners analysed in the population of study
(n = 731).
DFa
min
25th
50th
75th 90th
max
BDE-17
10.7 0.21 (0.20–0.22) 0.07
0.15
0.18
0.22 0.52
3.7
BDE-28
66.1 0.59 (0.55–0.63) 0.070
0.22
0.71
1.3
1.9
11.8
BDE-47
73.7
0.18
0.78
2.6
4.6
8.4
57.9
BDE-66
11.4 0.25 (0.24–0.26) 0.090
0.18
0.22
0.27 0.60
3.5
BDE-71
15.2 0.27 (0.26–0.29) 0.090
0.19
0.22
0.29 0.83
3.5
BDE-85
63.5 0.72 (0.66–0.80) 0.090
0.18
0.81
2.0
4.1
16.7
BDE-99
57.6
0.14
0.37
1.2
2.7
5.2
47.6
BDE-100
59.6 0.75 (0.69–0.80) 0.10
0.28
0.83
1.6
2.7
11.7
BDE-138
30.2 0.30 (0.28–0.32) 0.070
0.16
0.20
0.59
1.1
10.7
BDE-153
70.5 0.74 (0.68–0.80) 0.090
0.25
0.94
1.6
2.5
27.7
BDE-154
51.7 0.46 (0.43–0.49) 0.090
0.19
0.44
0.96
1.9
7.4
BDE-183
23.3 0.27 (0.26–0.29) 0.080
0.17
0.21
0.33 0.88
5.4
BDE-190
1.9
0.21 (0.20–0.21) 0.090
0.17
0.20
0.23 0.26
3.9
BDE-209
82.8
3.5
0.45
2.5
3.7
5.1
8.6
50.0
14.6 (13.9–15.4)
2.2
9.2
15.4
23.8 34.6
147
10.2
(9.7–10.8)
1.8
5.8
10.5
18.5 28.2
127
7.3
(6.9–7.8)
0.99
3.9
8.0
13.8 21.5
121
ΣPBDEsb
ΣPBDEsc
ΣPBDE3-7
715
716
717
718
a
d
GM
2.2
1.1
(95% CI)
(2.0–2.3)
(1.0–1.2)
(3.3–3.7)
Detection frequency (% of identification above limit of detection). bSum of all PBDE congeners
analysed. cSum of all PBDE congeners analysed except BDE-209. dSum of most commonly
reported
PBDE
congeners
(28,
47,
99,
100,
153,
154
and
183).
Table 3
Median PBDE concentrations (in ng/g lipid) found in human serum from different countries.
Location
Catalonia
Norway
Norway
Sweden
Sweden
UK
Belgium
Belgium
France
Greece
Faroe Islands
Spain
Spain
New Zealand
Australia
Korea
China (north)
China
China (south)
Japan
Japan
Canada
US
US
US
US
US
US
Year
2002
1998
1999
2000
2000
2003
1999
2004
2004
2007
1994
2003
2003
2001
2006
2001
2006
2007
2007
2003
1999
2000
2001
2003
2006
2007
n
731
69b
29b
15
50
154
11
8b
91
61
57
217
174
23
84b
62
115
12b
21
156
72
110
416
7b
168
2040
20
24
BDE-28 BDE-47 BDE-85
0.71
2.6
0.81
0.14
2.2
NA
0.24
1.5
NA
0.070
0.83 <0.010
NA
0.91
NA
<0.14
0.82
<0.13
0.10
1.2
ND
NA
0.97
NA
0.12
2.8
0.060
0.010
0.16
NA
NA
1.3
NA
0.060
2.5
0.16
ND
2.3
ND
NA
3.2
NA
g
8.4
0.30f
5.2
0.15
0.040
0.83
ND
29g
21g
NA
0.39
1.0
NA
0.14
0.74
0.071
0.080
0.32
0.010
h
NA
8.1
NA
0.40
15
ND
NA
34
0.70
0.013
0.11
0.013
1.1
19
ND
0.44
8.8
ND
1.3
17
<ND
BDE-99 BDE-100 BDE-153 BDE-154 BDE-183 BDE-209 ΣPBDEsa
1.2
0.83
0.94
0.44
0.21
3.7
15.4 (8.0)
0.45
0.45
0.54
0.26
ND
NA 4.0
0.31
0.35
0.59
0.35
ND
NA 3.3
0.19
0.17
0.56
0.040
0.060
NA 2.1
0.20
0.29
1.1
0.33
NA
0.46 (3.6)
<0.16
0.76
1.7
0.60
0.30
<15 5.6 (4.7d)
0.40
0.27
1.6
ND
0.21
11.1 3.8e
0.080
0.20
1.4
1.3
0.32
NA 4.6
1.9
0.37
0.72
0.065
0.21
5.8
(0.98e)
0.090
0.11
0.51
0.020
0.030
1.2
(1.1d)
f
0.33
0.51
1.0
1.4
NA
0.77 4.0
2.4
1.4
0.83
0.10
0.30
1.1
11.4 (8.5)
0.35
ND
2.1
1.5
ND
<0.7 9.6
0.80
0.62
1.0
0.080
0.23
NA (6.1)
g
g
g
2.5
2.0
3.3
16g
2.6
1.1
2.7
0.27
1.9
NA 16
0.67
0.13
0.32
0.040
0.41
ND 7.1 (2.9)
23g
15g
33g
42g
47g
403g 613g
0.36
0.15
1.4
0.10
0.31
NQ (4.4)
0.17
0.22
0.59
0.11
0.16
6.9
9.5
0.040
0.14
0.66
0.050
0.015
0.90 3.6
1.4h
1.1h
1.35h
NA
NA
NA (13.4h)
4.0
2.5
2.2
ND
ND
NA 27
11
5.9
7.3
0.95
NA
NA 61
0.032
0.013
0.025
NA
NA
NA 0.22
h
5.0
3.6
4.8
ND
ND
NA 33.5i
f
1.4
1.2
5.8
1.8
ND
1.4
17
2.4
3.0
7
ND
ND
ND 39
22
Reference
Present study
Thomsen et al. (2002)c
Thomsen et al. (2002)
Guvenius et al. (2003)
Weiss et al. (2006)
Thomas et al. (2006)
Covaci and Voorspoels (2005)
Roosens et al. (2010)
Antignac et al. (2009)
Kalantzi et al. (2011)
Fängström et al. (2005)
Gómara et al. (2007)c
Vizcaino et al. (2011)
Harrad and Porter (2007)
Toms et al. (2009)
Lee et al. (2007)
Zhu et al. (2009)
Jin et al. (2009)
Bi et al. (2006)
Takasuga et al. (2004)
Uemura et al. (2010)
Sandanger et al. (2007)
Castorina et al. (2011)
Sjödin et al. (2004)
Turyk et al. (2010)
Sjödin et al. (2008)
Lunder et al. (2010)
Johnson et al. (2010)
Sum of all congeners analysed in each study. Between brackets, sum of tri-hepta BDE congeners (∑PBDEs3−7). bPooled samples. cMedians of the whole population were not
available. They have been calculated from weighting of the medians of the different groups reported. dSum of BDEs 47, 99, 100, 153 and 154. eSum of all congeners except BDE209. fThe reported value coeluted with another compound. gMean instead of median. hGeometric mean instead of median. iSum of all geometric means of the congeners reported in
the
study.
NA:
Not
analysed;
ND:
Not
detected.
a
23
FIGURE CAPTIONS
Figure 1. Socio-demographic plots of the geometric means and 95% confidence intervals (ng/g
lipid) of the PBDE congeners found in highest concentrations. Age: 18-29 years, 30-44 years, 45-59
years, 60-74 years. Sex: women, men. Body mass index: normal, overweight, obesity. Place of
birth: Catalan, Spanish, other countries. Parity, 0, 1, 2, ≥3 children. Education: no formal education,
primary, secondary (1st stage), secondary (2nd stage), university. Social class: I (most affluent), II,
III, IV and V (least affluent).
Figure 2. Odds ratios and confidence intervals at 95% from multivariate regression models
performed with the PBDE congeners found in highest abundance.
Figure 3. Distribution of total PBDEs among different age groups.
24