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Concentrations of polybrominated diphenyl ethers in blood serum from New Zealand
Chemosphere 66 (2007) 2019–2023 www.elsevier.com/locate/chemosphere
Concentrations of polybrominated diphenyl ethers in blood serum from New Zealand Stuart Harrad a
a,*
, Lawrence Porter
b
Division of Environmental Health and Risk Management, Public Health Building, School of Geography, Earth, and Environmental Sciences, University of Birmingham, Birmingham B15 2TT, United Kingdom b AgriQuality Ltd, Wellington Laboratory, 1B Bell Road, Lower Hutt, New Zealand Received 6 March 2006; received in revised form 3 July 2006; accepted 13 July 2006 Available online 1 September 2006
Abstract Polybrominated diphenyl ethers (PBDEs) were measured in samples of human blood serumP taken from 23 donors in Wellington, New Zealand. Concentrations expressed as the sum of congeners 47, 99, 100, 153, 154, and 183 ( PBDE) were – at an average of 7.17 ng P PBDE g (lipid)1 – within the range reported for human tissues in Europe, but lower than in Australia and North America. The most likely source of this contamination is considered to be the release of PBDEs from imported consumer goods. The congenerPpattern observed is in line with that reported for human tissues outside North America, but shows a lower contribution of PBDE 47 to PBDE than observed in North Americans. NoP significant (p > 0.1) differences between concentrations in males and females were detected, and no relationship between donor age and PBDE concentration was observed. One donor displayed concentrations that were significantly elevated (i.e. > average +2 standard deviations) above those in others in this study. 2006 Elsevier Ltd. All rights reserved. Keywords: Australasia; Levels in humans; Brominated flame retardants
1. Introduction Polybrominated diphenyl ethers (PBDEs) are a group of brominated compounds widely used as flame retardants. In recent years, production and use of PBDEs has been in the guise of three formulations: penta (consisting primarily of BDEs 47 and 99–37% each, alongside smaller amounts of other tetra-, penta- and hexa-BDEs), octa [a mixture of hexa- (10–12%), hepta- (44–46%), octa- (33–35%), and nona-BDE (10–11%)], and deca (98% decabromodiphenyl ether – BDE 209 – and 2% various nona-BDEs) (McDonald, 2002; Alcock et al., 2003). Worldwide, PBDE production is dominated by the deca commercial formulation, with global demand in 2001 an estimated 56 100 t (BSEF, 2006). This is similar to the 1999 estimate of 54 800 t
*
Corresponding author. Tel.: +44 121 414 7298; fax: +44 121 414 3078. E-mail address: [email protected] (S. Harrad).
0045-6535/$ - see front matter 2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2006.07.042
(Renner, 2000). By comparison, 2001 global demand for the penta-product was 7500 t (BSEF, 2006), down slightly from 8500 t in 1999 (Renner, 2000), with the majority used in North America, for example, in 2001, 7100 t of pentaproduct were used in North America, compared to just 400 t elsewhere (BSEF, 2006). The uses for these commercial formulations are myriad. The penta-product was employed principally to flame retard polyurethane foams in carpet underlay, furniture and bedding. The octa-formulation was used to flame retard thermoplastics such as high impact polystyrene. The deca-product is used principally in plastic housings for electrical goods like TVs and computers, as well as textiles (Alcock et al., 2003). As a result of concerns surrounding these contaminants, owing to their presence in the diet and indoor air and dust (Harrad et al., 2004, 2006; Jones-Otazo et al., 2005; Stapleton et al., 2005), and human tissues (Hites, 2004), coupled with evidence relating to their potential adverse effects on human health (Darnerud et al., 2001; McDonald, 2002),
2020
S. Harrad, L. Porter / Chemosphere 66 (2007) 2019–2023
several jurisdictions (specifically the EU and several American states) have introduced restrictions on the marketing and use of penta- and octa-BDEs (BSEF, 2006). Furthermore, the main US producer has reached a voluntary agreement with the US EPA to discontinue production of the penta- and octa-BDE mixtures. Despite this, there remains a significant reservoir of PBDEs in use throughout the world. Concern has been expressed that, unless adequate strategies for end-of-life management for PBDEcontaining goods are developed, this reservoir will, in addition to providing a direct vector for both current and future exposure, ultimately escape into the environment where following incorporation into the food chain, it will result in elevated future exposure via the diet (Harrad and Diamond, 2006). As far as the authors are aware following extensive correspondence with New Zealand government officials, PBDEs have never been manufactured in New Zealand, nor imported directly as the commercial formulations. Furthermore, given the geographic isolation of New Zealand (the nearest major landmass – Australia – is around 2000 km distant, with southerly winds originating from the Antarctic), long range atmospheric/pelagic transport is unlikely to present a major source. However, it is highly likely that indirect importation of PBDEs into New Zealand will have occurred, in the guise of consumer goods (furnishings and electronic goods) treated in the country of manufacture with PBDE formulations. Despite this potential exposure of New Zealanders, coupled with a recent report that PBDE concentrations in Australian human milk exceed those in Europe (Harden et al., 2005), there are currently no data on the presence of these contaminants in the New Zealand environment, including human tissues. In 2001, as part of their method development work, AgriQuality conducted a survey of concentrations of POPs (including PBDEs) in blood samples from volunteers amongst their employees. Although the samples were analysed immediately, the data on PBDEs are only now reported in the light of burgeoning international interest in these compounds. In this paper our principal objectives were: • Make a preliminary assessment of the concentrations of these contaminants in New Zealanders. • Compare and contrast the concentrations and congener patterns with those found elsewhere in the world.
2. Experimental section 2.1. Sampling All samples were taken in 2001, and processed immediately. Each volunteer provided 100 ml of blood. A total of 23 individuals (10 male, 13 female, ages ranging between 20 and 64 years) provided samples, each of which was analysed separately.
2.2. Analytical protocols 2.2.1. Determination of serum lipid content In order to express concentration in samples on a lipid normalised basis, an accurately measured aliquot of sample was analysed for total lipid content. Total lipid (TL) was expressed in g l1, and calculated according to the algorithm recently reported by Covaci et al. (2006); viz TL = 1.33 · triglyceride + 1.12 · cholesterol + 1.48. 2.2.2. Determination of PBDEs Samples (62–100 ml serum, accurately measured) were diluted with distilled water (50 ml), and ethanol (75 ml), spiked with known quantities of internal standards (13C12-BDEs 47, 99, 153, and 183), shaken, and allowed to stand for 30 min. Following sequential extraction with diethyl ether:hexane (3 · 100 ml; 1:1 v/v), the hexane layers were combined and washed with distilled water (100 ml), and dried by passing through a microcolumn of anhydrous sodium sulfate. The hexane extract was then washed with concentrated sulfuric acid, before concentration to 1 ml and application (with 2 · 1 ml hexane rinses) to a microcolumn containing (from the top) silica (0.3 g) and sulfuric acid-impregnated silica (3 g; 44% w/w). The column was eluted with 9 ml hexane, with the eluate allowed to run directly onto the top of a second microcolumn packed with basic alumina (1.2 g). After elution of the first column, it was removed (and the entire eluate to that point discarded) and the second column alone eluted with dichloromethane:hexane (6 ml; 1:1 v/v). The entire eluate was concentrated to 0.5 ml before application (with 2 · 1 ml hexane rinses) to a microcolumn packed with CarbopackTM on celite (0.55 g) sandwiched between two small layers of anhydrous sodium sulfate. This column was eluted with hexane (8 ml), dichloromethane:cyclohexane (2 ml; 1:1 v/v), and dichloromethane:methanol:toluene (2 ml; 15:4:1 v/v). The entire eluate was combined and retained, prior to concentration and exchange of solvent to tetradecane (20 ll). GC/MS analysis was carried out on a Micromass Autospec Ultima mass spectrometer interfaced with an Agilent HP6890 GC fitted with an SGE ULTRA 2 column (30 m · 0.25 mm i.d., 0.1 lm film thickness). One ll of sample extract was injected in the splitless mode at an injector temperature of 300 C. The oven temperature programme was: 190 C for 1 min, 15 C/min to 300 C, held for 17 min. Sixteen ions (for BDE-47 and 13C12-BDE-47: 483.71, 485.71, 495.75, 497.75; BDEs 99 and 100, and 13 C12-BDE-99: 403.79, 405.79, 415.83, 417.83; BDE 153, 154, and 13C12-BDE-153: 481.70, 483.70, 493.74, 495.74; BDE 183 and 13C12-BDE-183: 561.61, 563.61, 573.65, 575.65) were monitored in 4 acquisition groups in EI selected ion monitoring mode (mass resolution = 5000; ionisation voltage, 40 eV; ion source temperature = 300 C). To ensure accurate and precise measurement, peaks were only accepted if the following criteria were met:
S. Harrad, L. Porter / Chemosphere 66 (2007) 2019–2023
• Signal to noise ratios for the least abundant ion exceeded 10:1. • Peaks eluted within 5 s of standards run in the same batch as the samples. • Isotope ratios for peaks were within 20% of those obtained for standards run in the same batch as the samples. Method blanks (n = 4) were analysed and found to contain concentrations of target PBDEs that were typically 10% of the concentrations found in the corresponding samples. Reporting limits were defined as that concentration equivalent to three times the average plus 3 times the standard deviation of concentrations detected in method blanks. Average recoveries of internal standards for all samples, ranged from 55% (13C12-BDE-99) to 95% (13C12BDE-47). The repeatability of the analytical procedures was evaluated by analysing ongoing precision and recovery (OPR) samples with each analytical batch spiked with analytes at the equivalent of 2 ng g1 lipid. The low relative standard deviations observed for concentrations of the target PBDE congeners (average = 5.2%; range 2.8–10.5%) demonstrate good repeatability for the analytical method. The accuracy of the method is indicated by the results obtained when AgriQuality participated in the 6th international interlaboratory trial on Dioxins and PCBs administered in 2005 by the Norwegian Institute for Public Health.
2021
Table 2 Concentrations (ng g1 lipid) of PBDEs in individual New Zealand blood serum samples Sample ID
Concentration of congener # 47
99
100
153
154
183
P BDEa
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23
2.38 1.94 12.7 1.69 <1.52 4.32 <1.52 3.23 8.9 <1.52 2.06 2.52 2.14 4.81 3.48 3.17 <1.52 4.81 3.93 8.06 3.77 3.13 6.23
0.72 0.57 1.81 0.46 0.43 2.34 0.41 0.88 2.09 0.36 0.91 1.01 0.63 0.92 0.50 1.24 0.32 0.90 0.90 1.49 0.68 0.54 1.51
0.54 0.41 2.47 0.33 0.30 1.75 0.34 0.68 1.31 <0.26 0.44 0.75 1.07 0.96 0.48 0.62 0.30 0.62 0.80 3.89 0.89 0.61 1.18
0.98 0.92 2.31 1.01 0.43 2.22 1.02 1.69 1.47 0.63 0.46 0.90 1.83 1.69 1.39 1.11 0.43 0.96 1.17 0.81 0.93 1.28 2.12
0.06 0.06 0.15 0.04 0.05 0.28 0.06 0.08 0.13 0.04 0.06 0.09 0.12 0.08 0.07 0.11 0.03 0.09 0.08 0.18 0.09 0.06 0.13
0.16 0.65 0.69 0.23 <0.12 0.17 0.90 0.14 0.50 0.22 0.16 0.35 0.17 0.15 0.37 0.29 0.12 0.22 0.24 0.97 0.30 0.20 0.68
4.84 4.54 20.1 3.76 2.85 11.1 4.25 6.69 14.4 3.03 4.09 5.62 5.95 8.62 6.29 6.53 2.72 7.60 7.13 15.4 6.66 5.82 11.9
a
Total lipid (g l1) 2.77 4.37 3.13 6.22 4.55 5.18 4.97 4.45 6.84 3.62 3.47 3.87 6.07 3.15 2.56 4.47 8.21 4.20 4.02 2.38 3.86 2.73 2.81
Upper bound value.
3. Results and discussion 3.1. PBDE concentrations in New Zealand serum Table 1 summarises the concentrations of target PBDEs in serum samples taken in this study, with concentrations in each individual sample provided in Table 2. The concentrations recorded in this study show that body burdens of PBDEs in New Zealanders are in line with those found in Europe and Japan, but lower than in North America (summarised in Hites, 2004). Most geographically relevant, is comparison with the recently reported concentrations in
pooled Australian human milk samples (n = 17), sampled in 2002–2003. This shows concentrations in this study to be lower but comparable to those in Australia (i.e. average sum of the same target BDEs is 10.04 ng g1 lipid) (Harden et al., 2005). While BDE 47 is the major congener detected, the second most prevalent congener is BDE 153 comprising 15–20% of all target BDEs. This congener pattern is again more akin to that detected outside North America, including that in Australian human milk (Harden et al., 2005). Given that as far as it has been possible to ascertain, there has never been any manufacture and direct use of 18
Table 1 Summary of concentrations (ng g1 lipid) of PBDEs in New Zealand blood serum Congener #
47 99 100 153 154 183 P PBDE a
Upper bounda
Lower boundb
16 14 12
Average
rn1
Median
Average
rn1
Median
3.79 0.91 0.89 1.21 0.09 0.34 7.17
2.77 0.57 0.83 0.55 0.06 0.25 4.45
3.15 0.80 0.62 1.02 0.08 0.23 6.12
3.47 0.91 0.86 1.16 0.09 0.33 6.83
3.09 0.57 0.85 0.59 0.06 0.26 4.83
3.15 0.80 0.62 1.01 0.08 0.23 6.12
Upper bound (i.e. where the concentration of a congener was below the detection limit, it was assumed to equal the detection limit). b Lower bound (i.e. where the concentration of a congener was below the detection limit, it was assumed to be zero).
10 8 6 4 2 0
20-30
31-40
41-50
51-64
Donor Age (Years) Fig. 1. Relationship between donor age and concentration of (error bars are ±1r).
P PBDE
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S. Harrad, L. Porter / Chemosphere 66 (2007) 2019–2023 8 7
Frequency (n)
6 5 4 3 2 1 0 0.0-2.0 2.1-4.0 4.1-6.0 6.1-8.0 8.1-10.0
10.112.0
12.114.0
14.116.0
16.118.0
18.120.0
20.122.0
Concentration (ng ΣBDE g-1) Fig. 2. Frequency distribution of concentrations of
PBDEs in New Zealand, there are two likely pathways of exposure. One is via PBDEs present in imported foodstuffs; the other is exposure to PBDEs present in imported household and consumer goods. At the present time, in the absence of any data on concentrations of PBDEs in the New Zealand diet, and in indoor air and dusts – each of which have been identified as significant sources of human exposure (Harrad et al., 2004, 2006; Jones-Otazo et al., 2005; Stapleton et al., 2005) – it is not possible to definitively determine the relative significance of these exposure pathways for the New Zealand population. 3.2. Relationship between donor age and PBDE concentrations Previous studies have reported that in contrast to dioxin-like compounds, for which concentrations in human tissues increase with donor age, there is no relationship between donor age and concentration of PBDEs (Hardell et al., 1998; Sjo¨din et al., 2000; Covaci et al., P 2002). Fig. 1 compares concentrations of PBDE in 4 age groups: 20–30, 31–40, 41–50, 51–64. No significant relationship is apparent. 3.3. Male:female concentration differences Although average concentrationsP in male donors were higher than in females (8.32 ng BDE g1 lipid c.f. 6.67), a t-test revealed the difference to be statistically insignificant (p > 0.1). 3.4. Individual variability in concentrations P Fig. 2 shows a frequency distribution histogram of PBDE concentrations found in individual donors. It illustrates that the data are not normally distributed. Using a definition of an outlier as a concentration that exceeds the average +2 standard P deviations, one individual can be defined as having PBDE concentrations that are sig-
P
PBDE.
nificantly above those detected in the majority of samples in this study. This feature of our dataset is consistent with previous reports that a small minority of individuals display body burdens that are significantly elevated (van Bavel et al., 2002). While we have no data on the potential exposure patterns of individual donors that would facilitate identification of the causes of such elevated body burdens, other authors (Harrad and Diamond, 2006) have hypothesised that the heterogeneity of indoor exposures (Harrad et al., 2006) may provide the explanation. This study provides the first evidence that the New Zealand population is contaminated with PBDEs at concentrations in line with those in Europe. It suggests that despite New Zealand never having manufactured these chemicals, their widespread incorporation into a wide range of imported consumer goods, has led to discernible exposure of New Zealanders. Furthermore, the detection in a geographically remote location of a class of chemicals that is being considered for future inclusion within the scope of the Stockholm convention on POPs; points to the fact that it is not simply their environmental persistence that leads to their worldwide distribution far from the points of manufacture, but international trade in treated manufactured goods. Investigations are urgently needed into the pathways via which this exposure to New Zealanders occurs, and its toxicological significance.
Acknowledgements The authors gratefully acknowledge the provision of a Short Visit Grant to Stuart Harrad from the Royal Society.
References Alcock, R.E., Sweetman, A.J., Prevedouros, K., Jones, K.C., 2003. Understanding levels and trends of BDE-47 in the UK and North America: an assessment of principal reservoirs and source inputs. Environ. Int. 29, 691–698.
S. Harrad, L. Porter / Chemosphere 66 (2007) 2019–2023 BSEF (Bromine Science Environmental Forum), 2006. (accessed July). 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. A 88, 210–218. Covaci, A., Voorspoels, S., Thomsen, C., van Bavel, B., Neels, H., 2006. Evaluation of total lipids using enzymatic methods for the normalization of persistent organic pollutant levels in serum. Sci. Total Environ. 366, 361–366. Darnerud, P.O., Eriksen, G.S., Jo´hannesson, T., Larsen, P.B., Viluksela, M., 2001. Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environ. Health Perspect. 109 (Suppl. 1), 49–68. Hardell, L., Linstrom, G., van Bavel, B., Wingfors, H., Sundelin, E., Liljegren, G., 1998. Concentrations of 2,2 0 ,4,4 0 -tetrabrominated diphenyl ether in human adipose tissue in Swedish persons and the risk for non-Hodgkin’s lymphoma. Oncol. Res. 10, 429–432. Harden, F., Mu¨ller, J., Toms, L., 2005. Organochlorine Pesticides (OCPs) and Polybrominated Diphenyl Ethers (PBDEs) in the Australian Population: Levels in Human Milk. Environment Protection and Heritage Council of Australia and New Zealand. ISBN 0 642 32384 4. Harrad, S., Diamond, M., 2006. Exposure to PBDEs and PCBs: current and future scenarios. Atmos. Environ. 40, 1187–1188. Harrad, S., Wijesekera, R., Hunter, S., Halliwell, C., Baker, R.A., 2004. Preliminary assessment of UK human dietary and inhalation exposure to polybrominated diphenyl ethers. Environ. Sci. Technol. 38, 2345– 2350.
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Harrad, S., Hazrati, S., Ibarra, C., 2006. 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. Hites, R.A., 2004. Polybrominated diphenyl ethers in the environment and in people: a meta-analysis of concentrations. Environ. Sci. Technol. 38, 945–956. Jones-Otazo, H., Clarke, J.P., Diamond, M.L., Archbold, J.A., Ferguson, G., Harner, T., Richardson, G.M., Ryan, J.J., Wilford, B., 2005. Is house dust the missing exposure pathway for PBDEs? An analysis of the urban fate and human exposure to PBDEs. Environ. Sci. Technol. 39, 5121–5130. McDonald, T.A., 2002. A perspective on the potential health risks of PBDEs. Chemosphere 46, 745–755. Renner, R., 2000. Increasing levels of flame retardants found in North American environment. Environ. Sci. Technol. 34, 452A–453A. Sjo¨din, A., Hagmar, L., Klasson-Wehler, E., Bjork, L., 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, 1035–1048. 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. van Bavel, B., Hardell, L., Kitti, A., Liljegadhl, A., Karlsson, M., Petersson, A., Tysklind, M., Lindstro¨m, G., 2002. Organohalogen Compd. 58, 161–164.
Concentrations of polybrominated diphenyl ethers in blood serum from New Zealand Stuart Harrad a
a,*
, Lawrence Porter
b
Division of Environmental Health and Risk Management, Public Health Building, School of Geography, Earth, and Environmental Sciences, University of Birmingham, Birmingham B15 2TT, United Kingdom b AgriQuality Ltd, Wellington Laboratory, 1B Bell Road, Lower Hutt, New Zealand Received 6 March 2006; received in revised form 3 July 2006; accepted 13 July 2006 Available online 1 September 2006
Abstract Polybrominated diphenyl ethers (PBDEs) were measured in samples of human blood serumP taken from 23 donors in Wellington, New Zealand. Concentrations expressed as the sum of congeners 47, 99, 100, 153, 154, and 183 ( PBDE) were – at an average of 7.17 ng P PBDE g (lipid)1 – within the range reported for human tissues in Europe, but lower than in Australia and North America. The most likely source of this contamination is considered to be the release of PBDEs from imported consumer goods. The congenerPpattern observed is in line with that reported for human tissues outside North America, but shows a lower contribution of PBDE 47 to PBDE than observed in North Americans. NoP significant (p > 0.1) differences between concentrations in males and females were detected, and no relationship between donor age and PBDE concentration was observed. One donor displayed concentrations that were significantly elevated (i.e. > average +2 standard deviations) above those in others in this study. 2006 Elsevier Ltd. All rights reserved. Keywords: Australasia; Levels in humans; Brominated flame retardants
1. Introduction Polybrominated diphenyl ethers (PBDEs) are a group of brominated compounds widely used as flame retardants. In recent years, production and use of PBDEs has been in the guise of three formulations: penta (consisting primarily of BDEs 47 and 99–37% each, alongside smaller amounts of other tetra-, penta- and hexa-BDEs), octa [a mixture of hexa- (10–12%), hepta- (44–46%), octa- (33–35%), and nona-BDE (10–11%)], and deca (98% decabromodiphenyl ether – BDE 209 – and 2% various nona-BDEs) (McDonald, 2002; Alcock et al., 2003). Worldwide, PBDE production is dominated by the deca commercial formulation, with global demand in 2001 an estimated 56 100 t (BSEF, 2006). This is similar to the 1999 estimate of 54 800 t
*
Corresponding author. Tel.: +44 121 414 7298; fax: +44 121 414 3078. E-mail address: [email protected] (S. Harrad).
0045-6535/$ - see front matter 2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2006.07.042
(Renner, 2000). By comparison, 2001 global demand for the penta-product was 7500 t (BSEF, 2006), down slightly from 8500 t in 1999 (Renner, 2000), with the majority used in North America, for example, in 2001, 7100 t of pentaproduct were used in North America, compared to just 400 t elsewhere (BSEF, 2006). The uses for these commercial formulations are myriad. The penta-product was employed principally to flame retard polyurethane foams in carpet underlay, furniture and bedding. The octa-formulation was used to flame retard thermoplastics such as high impact polystyrene. The deca-product is used principally in plastic housings for electrical goods like TVs and computers, as well as textiles (Alcock et al., 2003). As a result of concerns surrounding these contaminants, owing to their presence in the diet and indoor air and dust (Harrad et al., 2004, 2006; Jones-Otazo et al., 2005; Stapleton et al., 2005), and human tissues (Hites, 2004), coupled with evidence relating to their potential adverse effects on human health (Darnerud et al., 2001; McDonald, 2002),
2020
S. Harrad, L. Porter / Chemosphere 66 (2007) 2019–2023
several jurisdictions (specifically the EU and several American states) have introduced restrictions on the marketing and use of penta- and octa-BDEs (BSEF, 2006). Furthermore, the main US producer has reached a voluntary agreement with the US EPA to discontinue production of the penta- and octa-BDE mixtures. Despite this, there remains a significant reservoir of PBDEs in use throughout the world. Concern has been expressed that, unless adequate strategies for end-of-life management for PBDEcontaining goods are developed, this reservoir will, in addition to providing a direct vector for both current and future exposure, ultimately escape into the environment where following incorporation into the food chain, it will result in elevated future exposure via the diet (Harrad and Diamond, 2006). As far as the authors are aware following extensive correspondence with New Zealand government officials, PBDEs have never been manufactured in New Zealand, nor imported directly as the commercial formulations. Furthermore, given the geographic isolation of New Zealand (the nearest major landmass – Australia – is around 2000 km distant, with southerly winds originating from the Antarctic), long range atmospheric/pelagic transport is unlikely to present a major source. However, it is highly likely that indirect importation of PBDEs into New Zealand will have occurred, in the guise of consumer goods (furnishings and electronic goods) treated in the country of manufacture with PBDE formulations. Despite this potential exposure of New Zealanders, coupled with a recent report that PBDE concentrations in Australian human milk exceed those in Europe (Harden et al., 2005), there are currently no data on the presence of these contaminants in the New Zealand environment, including human tissues. In 2001, as part of their method development work, AgriQuality conducted a survey of concentrations of POPs (including PBDEs) in blood samples from volunteers amongst their employees. Although the samples were analysed immediately, the data on PBDEs are only now reported in the light of burgeoning international interest in these compounds. In this paper our principal objectives were: • Make a preliminary assessment of the concentrations of these contaminants in New Zealanders. • Compare and contrast the concentrations and congener patterns with those found elsewhere in the world.
2. Experimental section 2.1. Sampling All samples were taken in 2001, and processed immediately. Each volunteer provided 100 ml of blood. A total of 23 individuals (10 male, 13 female, ages ranging between 20 and 64 years) provided samples, each of which was analysed separately.
2.2. Analytical protocols 2.2.1. Determination of serum lipid content In order to express concentration in samples on a lipid normalised basis, an accurately measured aliquot of sample was analysed for total lipid content. Total lipid (TL) was expressed in g l1, and calculated according to the algorithm recently reported by Covaci et al. (2006); viz TL = 1.33 · triglyceride + 1.12 · cholesterol + 1.48. 2.2.2. Determination of PBDEs Samples (62–100 ml serum, accurately measured) were diluted with distilled water (50 ml), and ethanol (75 ml), spiked with known quantities of internal standards (13C12-BDEs 47, 99, 153, and 183), shaken, and allowed to stand for 30 min. Following sequential extraction with diethyl ether:hexane (3 · 100 ml; 1:1 v/v), the hexane layers were combined and washed with distilled water (100 ml), and dried by passing through a microcolumn of anhydrous sodium sulfate. The hexane extract was then washed with concentrated sulfuric acid, before concentration to 1 ml and application (with 2 · 1 ml hexane rinses) to a microcolumn containing (from the top) silica (0.3 g) and sulfuric acid-impregnated silica (3 g; 44% w/w). The column was eluted with 9 ml hexane, with the eluate allowed to run directly onto the top of a second microcolumn packed with basic alumina (1.2 g). After elution of the first column, it was removed (and the entire eluate to that point discarded) and the second column alone eluted with dichloromethane:hexane (6 ml; 1:1 v/v). The entire eluate was concentrated to 0.5 ml before application (with 2 · 1 ml hexane rinses) to a microcolumn packed with CarbopackTM on celite (0.55 g) sandwiched between two small layers of anhydrous sodium sulfate. This column was eluted with hexane (8 ml), dichloromethane:cyclohexane (2 ml; 1:1 v/v), and dichloromethane:methanol:toluene (2 ml; 15:4:1 v/v). The entire eluate was combined and retained, prior to concentration and exchange of solvent to tetradecane (20 ll). GC/MS analysis was carried out on a Micromass Autospec Ultima mass spectrometer interfaced with an Agilent HP6890 GC fitted with an SGE ULTRA 2 column (30 m · 0.25 mm i.d., 0.1 lm film thickness). One ll of sample extract was injected in the splitless mode at an injector temperature of 300 C. The oven temperature programme was: 190 C for 1 min, 15 C/min to 300 C, held for 17 min. Sixteen ions (for BDE-47 and 13C12-BDE-47: 483.71, 485.71, 495.75, 497.75; BDEs 99 and 100, and 13 C12-BDE-99: 403.79, 405.79, 415.83, 417.83; BDE 153, 154, and 13C12-BDE-153: 481.70, 483.70, 493.74, 495.74; BDE 183 and 13C12-BDE-183: 561.61, 563.61, 573.65, 575.65) were monitored in 4 acquisition groups in EI selected ion monitoring mode (mass resolution = 5000; ionisation voltage, 40 eV; ion source temperature = 300 C). To ensure accurate and precise measurement, peaks were only accepted if the following criteria were met:
S. Harrad, L. Porter / Chemosphere 66 (2007) 2019–2023
• Signal to noise ratios for the least abundant ion exceeded 10:1. • Peaks eluted within 5 s of standards run in the same batch as the samples. • Isotope ratios for peaks were within 20% of those obtained for standards run in the same batch as the samples. Method blanks (n = 4) were analysed and found to contain concentrations of target PBDEs that were typically 10% of the concentrations found in the corresponding samples. Reporting limits were defined as that concentration equivalent to three times the average plus 3 times the standard deviation of concentrations detected in method blanks. Average recoveries of internal standards for all samples, ranged from 55% (13C12-BDE-99) to 95% (13C12BDE-47). The repeatability of the analytical procedures was evaluated by analysing ongoing precision and recovery (OPR) samples with each analytical batch spiked with analytes at the equivalent of 2 ng g1 lipid. The low relative standard deviations observed for concentrations of the target PBDE congeners (average = 5.2%; range 2.8–10.5%) demonstrate good repeatability for the analytical method. The accuracy of the method is indicated by the results obtained when AgriQuality participated in the 6th international interlaboratory trial on Dioxins and PCBs administered in 2005 by the Norwegian Institute for Public Health.
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Table 2 Concentrations (ng g1 lipid) of PBDEs in individual New Zealand blood serum samples Sample ID
Concentration of congener # 47
99
100
153
154
183
P BDEa
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23
2.38 1.94 12.7 1.69 <1.52 4.32 <1.52 3.23 8.9 <1.52 2.06 2.52 2.14 4.81 3.48 3.17 <1.52 4.81 3.93 8.06 3.77 3.13 6.23
0.72 0.57 1.81 0.46 0.43 2.34 0.41 0.88 2.09 0.36 0.91 1.01 0.63 0.92 0.50 1.24 0.32 0.90 0.90 1.49 0.68 0.54 1.51
0.54 0.41 2.47 0.33 0.30 1.75 0.34 0.68 1.31 <0.26 0.44 0.75 1.07 0.96 0.48 0.62 0.30 0.62 0.80 3.89 0.89 0.61 1.18
0.98 0.92 2.31 1.01 0.43 2.22 1.02 1.69 1.47 0.63 0.46 0.90 1.83 1.69 1.39 1.11 0.43 0.96 1.17 0.81 0.93 1.28 2.12
0.06 0.06 0.15 0.04 0.05 0.28 0.06 0.08 0.13 0.04 0.06 0.09 0.12 0.08 0.07 0.11 0.03 0.09 0.08 0.18 0.09 0.06 0.13
0.16 0.65 0.69 0.23 <0.12 0.17 0.90 0.14 0.50 0.22 0.16 0.35 0.17 0.15 0.37 0.29 0.12 0.22 0.24 0.97 0.30 0.20 0.68
4.84 4.54 20.1 3.76 2.85 11.1 4.25 6.69 14.4 3.03 4.09 5.62 5.95 8.62 6.29 6.53 2.72 7.60 7.13 15.4 6.66 5.82 11.9
a
Total lipid (g l1) 2.77 4.37 3.13 6.22 4.55 5.18 4.97 4.45 6.84 3.62 3.47 3.87 6.07 3.15 2.56 4.47 8.21 4.20 4.02 2.38 3.86 2.73 2.81
Upper bound value.
3. Results and discussion 3.1. PBDE concentrations in New Zealand serum Table 1 summarises the concentrations of target PBDEs in serum samples taken in this study, with concentrations in each individual sample provided in Table 2. The concentrations recorded in this study show that body burdens of PBDEs in New Zealanders are in line with those found in Europe and Japan, but lower than in North America (summarised in Hites, 2004). Most geographically relevant, is comparison with the recently reported concentrations in
pooled Australian human milk samples (n = 17), sampled in 2002–2003. This shows concentrations in this study to be lower but comparable to those in Australia (i.e. average sum of the same target BDEs is 10.04 ng g1 lipid) (Harden et al., 2005). While BDE 47 is the major congener detected, the second most prevalent congener is BDE 153 comprising 15–20% of all target BDEs. This congener pattern is again more akin to that detected outside North America, including that in Australian human milk (Harden et al., 2005). Given that as far as it has been possible to ascertain, there has never been any manufacture and direct use of 18
Table 1 Summary of concentrations (ng g1 lipid) of PBDEs in New Zealand blood serum Congener #
47 99 100 153 154 183 P PBDE a
Upper bounda
Lower boundb
16 14 12
Average
rn1
Median
Average
rn1
Median
3.79 0.91 0.89 1.21 0.09 0.34 7.17
2.77 0.57 0.83 0.55 0.06 0.25 4.45
3.15 0.80 0.62 1.02 0.08 0.23 6.12
3.47 0.91 0.86 1.16 0.09 0.33 6.83
3.09 0.57 0.85 0.59 0.06 0.26 4.83
3.15 0.80 0.62 1.01 0.08 0.23 6.12
Upper bound (i.e. where the concentration of a congener was below the detection limit, it was assumed to equal the detection limit). b Lower bound (i.e. where the concentration of a congener was below the detection limit, it was assumed to be zero).
10 8 6 4 2 0
20-30
31-40
41-50
51-64
Donor Age (Years) Fig. 1. Relationship between donor age and concentration of (error bars are ±1r).
P PBDE
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S. Harrad, L. Porter / Chemosphere 66 (2007) 2019–2023 8 7
Frequency (n)
6 5 4 3 2 1 0 0.0-2.0 2.1-4.0 4.1-6.0 6.1-8.0 8.1-10.0
10.112.0
12.114.0
14.116.0
16.118.0
18.120.0
20.122.0
Concentration (ng ΣBDE g-1) Fig. 2. Frequency distribution of concentrations of
PBDEs in New Zealand, there are two likely pathways of exposure. One is via PBDEs present in imported foodstuffs; the other is exposure to PBDEs present in imported household and consumer goods. At the present time, in the absence of any data on concentrations of PBDEs in the New Zealand diet, and in indoor air and dusts – each of which have been identified as significant sources of human exposure (Harrad et al., 2004, 2006; Jones-Otazo et al., 2005; Stapleton et al., 2005) – it is not possible to definitively determine the relative significance of these exposure pathways for the New Zealand population. 3.2. Relationship between donor age and PBDE concentrations Previous studies have reported that in contrast to dioxin-like compounds, for which concentrations in human tissues increase with donor age, there is no relationship between donor age and concentration of PBDEs (Hardell et al., 1998; Sjo¨din et al., 2000; Covaci et al., P 2002). Fig. 1 compares concentrations of PBDE in 4 age groups: 20–30, 31–40, 41–50, 51–64. No significant relationship is apparent. 3.3. Male:female concentration differences Although average concentrationsP in male donors were higher than in females (8.32 ng BDE g1 lipid c.f. 6.67), a t-test revealed the difference to be statistically insignificant (p > 0.1). 3.4. Individual variability in concentrations P Fig. 2 shows a frequency distribution histogram of PBDE concentrations found in individual donors. It illustrates that the data are not normally distributed. Using a definition of an outlier as a concentration that exceeds the average +2 standard P deviations, one individual can be defined as having PBDE concentrations that are sig-
P
PBDE.
nificantly above those detected in the majority of samples in this study. This feature of our dataset is consistent with previous reports that a small minority of individuals display body burdens that are significantly elevated (van Bavel et al., 2002). While we have no data on the potential exposure patterns of individual donors that would facilitate identification of the causes of such elevated body burdens, other authors (Harrad and Diamond, 2006) have hypothesised that the heterogeneity of indoor exposures (Harrad et al., 2006) may provide the explanation. This study provides the first evidence that the New Zealand population is contaminated with PBDEs at concentrations in line with those in Europe. It suggests that despite New Zealand never having manufactured these chemicals, their widespread incorporation into a wide range of imported consumer goods, has led to discernible exposure of New Zealanders. Furthermore, the detection in a geographically remote location of a class of chemicals that is being considered for future inclusion within the scope of the Stockholm convention on POPs; points to the fact that it is not simply their environmental persistence that leads to their worldwide distribution far from the points of manufacture, but international trade in treated manufactured goods. Investigations are urgently needed into the pathways via which this exposure to New Zealanders occurs, and its toxicological significance.
Acknowledgements The authors gratefully acknowledge the provision of a Short Visit Grant to Stuart Harrad from the Royal Society.
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