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Levels and chiral signatures of persistent organochlorine pollutants in human tissues from Belgium
ARTICLE IN PRESS
Environmental Research 93 (2003) 167–176
Levels and chiral signatures of persistent organochlorine pollutants in human tissues from Belgium Shaogang Chu, Adrian Covaci, and Paul Schepens Toxicological Center, University of Antwerp, Universiteitsplein 1, Antwerpen (Wilrijk) 2610, Belgium Received 8 August 2002; received in revised form 7 January 2003; accepted 17 January 2003
Abstract Polychlorinated biphenyls (PCBs) and organochlorine pesticides (OCPs) were measured in human tissue samples including muscle, liver, brain, and kidney. The samples were obtained at autopsy in 2000–2001 from three women and eight men from Belgium, aged between 5 and 76 years. The measured PCBs included 23 ortho-substituted congeners and 3 non-ortho-substituted congeners (PCB 77, PCB 126, and PCB 169). The mean concentrations of SPCBs were 29.4, 35.3, 10.6, and 11.8 ng/g wet wt in liver, muscle, kidney, and brain, respectively. HCB, g-HCH, b-HCH, p; p0 -DDE, and p; p0 -DDD were found in all samples, while p; p0 DDT could only be found in one liver sample. The most abundant pesticide was p; p0 -DDE. PCB 153 and PCB 180 were the main ortho-substituted congeners found in all the samples, while the concentration of the congeners with less than three chlorine atoms was below the limit of determination. In 10 of 18 tissues, the concentrations of PCB 169 were higher than the concentration of PCB 126. These results are consistent with the order of half-life of these congeners in humans and indicate that a steady state had been reached in these subjects. The enantiomeric compositions of a-HCH and chiral PCBs, including PCB 95, PCB 132, and PCB 149, were also measured. a-HCH was found to be racemic in three liver samples, while chiral PCB 95, PCB 149, and 132 showed racemic or nearly racemic compositions in muscle, kidney, and brain. Higher enatiomeric ratios (ERs) for the three chiral PCBs were found in liver samples. The mean (range) ERs in liver were 1.69 (1.04–2.97), 1.16 (0.99–1.41), and 0.74 (0.48–0.97) for PCB 95, PCB 149, and PCB 132, respectively. r 2003 Elsevier Science (USA). All rights reserved.
1. Introduction Polychlorinated biphenyls (PCBs) and organochlorine pesticides (OCPs) are ubiquitous pollutants that can be biologically amplified in food webs to high concentrations. Despite restrictions on their manufacture and use in most industrialized countries since the 1970s, they are still of public concern and accidents may occasionally occur (Chen et al., 1980; Kuratsune et al., 1972). In January 1999, 500 tons of feed contaminated with approximately 50 kg of PCBs and 1 g of dioxins were distributed to animal farms in Belgium (van Larebeke et al., 2001). An extensive investigation of possible effects caused by the incident has been done in the last 2 years (van Larebeke et al., 2001). Significant advancements have been introduced in PCB analysis, which directly impacted the validity of reported trend analyses. In the last decade, more
Corresponding author. Fax: +32-38202722. E-mail address: [email protected] (A. Covaci).
attention has been paid to the determination of nonortho PCBs, which show similar toxicity to polychlorinated dibenzo-p-dioxins and dibenzofurans (Schecter, 1998). More recently, attention has been focused on PCBs and other organochlorine contaminants that display axial chirality, and it was suggested that the enantiomeric ratio of chiral contaminants in different species may give additional information on possible bioaccumulation and degradation pathways (Kallenborn and Huhnerfuss, 2001; Wong et al., 2001). While there is now evidence that PCB concentrations in the environment were generally declining last decade, data available on humans suggest that the declining was much slower to respond to the restriction in use or incident. Presumably, this is due to the high persistence of PCBs in the fatty tissues in which they are deposited (Erickson, 1997). The data on human contamination with PCB and pesticides are focused on blood or milk, the only biological materials conveniently available from healthy subjects. In most cases, blood reflects the event of actual pollution exposure (Phillips et al., 1989;
0013-9351/03/$ - see front matter r 2003 Elsevier Science (USA). All rights reserved. doi:10.1016/S0013-9351(03)00016-1
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Schecter, 1998) and selective metabolism and excretion may alter the profile. Adipose tissue is another choice for the investigation of human body burden (Phillips and Birchard, 1991). However, chemical behavior and biomagnification may vary among organs of the studied biota (Boon et al., 1987). Because human organs are not convenient to obtain, relatively limited data on the levels of PCBs and in particular of non-ortho PCBs in human tissues (other than adipose tissue) are now available. Moreover, data on the enantiomeric composition of PCB atropisomers in human tissues have not been reported at this time. The aim of this study was to investigate the levels and distribution of PCBs (including non-ortho congeners) and OCPs in selected human tissues from the Belgian general population and to explore eventual relationships with the enantiomeric ratios (ERs) of chiral pollutants.
Table 1 Identification of human tissue samples from Belgium
2. Materials and methods 2.1. Samples The sample materials consisted of postmortem human tissues from 11 deceased males and females (age range 5–76 years) that died suddenly from causes unrelated to environmental contaminants. All samples were collected in 2000–2001 at one Belgian hospital. Parts of the brain, liver, kidney, and muscle were obtained from unembalmed cadavers during autopsies and frozen at 201C until analysis. Sample details are listed in Table 1. Subjects 4–6 came from the same family. 2.2. Chemicals and standards All solvents were of pesticide residue analysis from Merck (Wessel, Germany). Individual PCB congeners and pesticide standards were purchased from Dr. Ehrenstorfer (Augsburg, Germany), while 13C12-labeled non-ortho-substituted PCBs (PCB 77, 126, 169) were obtained from Cambridge Isotope Laboratories (Andover, MA, USA). 2.3. Extraction, clean-up and fractionation Eight grams of tissue samples was ground with sufficient Na2SO4 to disintegrate the sample and bind the water. Internal standards (PCB 143, PCB 46, and eHCH for the determination of ortho-substituted PCBs and OCPs; 13C12-labeled PCB 77, 126, and 169 for the determination of non-ortho PCBs) were added to each tissue and blank sample prior to extraction. The samples were extracted for 4 h with hexane/acetone (3:1, v/v) in a BUCHI B-811 Soxhlet extraction system (Bruxelles, Belgium) operated in hot extraction mode. After concentration to approximately 5 mL, an aliquot
Subject no.
Sex
Age
Sample no.
Tissue
Lipid content (%)
1
Male
Adult
1B 1K 1L 1M
Brain Kidney Liver Muscle
0.4 2.9 3.6 3.7
2
Male
36
2K 2L 2M
Kidney Liver Muscle
1.8 5.8 1.4
3
Male Male Male
Adult
3K 3L 3M
Kidney Liver Muscle
3.6 5.8 3.1
4 5 6
Male Female Male
35 31 5
4L 5L 6L
Liver Liver Liver
3.6 4.4 5.5
7 8 9 10 11
Male Female Male Female Male
Child Child Adult 56 76
7L 8L 9L 10L 11L
Liver Liver Liver Liver Liver
3.8 2.6 13.8 3.7 5.7
of the extract was used to determine the extractable lipids gravimetrically. The remaining extract was cleaned on a column packed with 8 g acid silica (40% sulfuric acid, w/w), followed by a column with 10 g Florisil. One fraction (70 mL hexane) was collected and concentrated to 50 mL. The fractionation of PCB congeners according to their chlorine substitution was achieved by the method reported by Ramos et al. (1999) with some modifications. PCBs were separated by a 250 mm 4.6-mm i.d. Cosmosil 5-PYE column (Nacalai Tesque, Osaka, Japan). The HPLC system consisted of a Gilson 307 pump, 112 UV/VIS detector (Middleton, WI, USA), and a Rheodyne 7125 injector (Rohnert Park, CA, USA) equipped with a 20-mL sample loop. The detector was operated at 254 nm. Hexane was used as a mobile phase at a flow rate of 0.5 mL/min and the separated fractions were collected by a HeliRac LKB2212 fraction collector (LKB-Produkter AB, Bromma, Sweden). The first fraction (6–12 min) contained most ortho-substituted congeners and pesticides, while the second fraction (12–20 min) contained non-ortho-substituted congeners (PCB 77, PCB 126, PCB 169) and mono-ortho-substituted congener PCB 189. For the determination of the enantiomeric composition of PCB atropisomers, the same HPLC system was used to separate the individual chiral PCBs from the other congeners. Heart cut technology was used and fractions of 80 mL were collected for each chiral PCB (PCB 84, PCB 95, PCB 132, and PCB 149) according to their elution times on the HPLC chromatogram.
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2.4. Analysis An Agilent 6890 GC with microcell ECD was used to determine PCBs and pesticides. The separation was performed on a 25 m 0.22 mm, 0.25-mm-film-thickness HT-8 capillary column from SGE (Zulte, Belgium). The carrier and make-up gas were helium and argon/ methane (95:5), respectively. The initial oven temperature was 901C, hold for 1 min, and then to 1801C at 151C/min, hold for 1 min, to 2501C at 31C/min, and to 2901C at 151C and hold for 6 min. The injector temperature and detector temperature were 2701C and 3201C, respectively. A 1-mL extract was injected in pulse splitless mode and the purge time was 1 min. The identification was based on the retention time of individual standard congeners and a multilevel calibration curve was used for quantitative determination. For confirmation, all samples were reanalyzed by GC/ MS with an electron capture negative-impact ionization (ECNI) source and selected ion monitoring mode. An Agilent 6890 GC coupled with Agilent 5973N MSD was used for qualitative and quantitative analyses. A 1-mL aliquot of extract was injected into a 10 m 0.1 mm, 0.10-mm-thickness HT-8 (SGE) narrow-bore capillary column in solvent vent injection mode. The injection parameters were: initial temperature 601C, hold for 1 min, and then to 3001C at 6001C/min and hold for 20 min. The vent time was 0.9 min. The initial oven temperature was 601C, hold for 3 min, and then to 1501C at 501C/min, to 2501C at 101C/min, and to 2901C/min at 201C/min and hold for 10 min. The ion source and quadrupole temperatures were 2001C and 1501C, respectively. Methane was used as the reagent gas. Three characteristic ions of each analyte (tetra- to octa-CB and pesticides) were selected with seven retention windows. Identification was based on retention times and the intensity of the acquired ions in sample peaks within 10% of the mean values obtained from the corresponding standards. Quantitative determination was done using five-level calibration curves obtained from a standard solution containing all congeners. The method used for the analysis of non-ortho PCBs was similar to that described above, except that the injector system was operated in large volume injection mode. A 5 5-mL aliquot extract was injected into injector and the purge time was 2.2 min. Procedural blanks were subjected to the entire analytical procedure to determine background interference, and peaks were quantified only when the signal/ noise ratio was 43. The mean recovery of internal standards in individual samples was 75.1%. The limits of determination varied for individual PCB congeners and detector, and these generally decreased with increased levels of chlorination. Limits of determination for ortho-substituted PCBs and pesticides ranged from
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0.01 ng/g for penta-CB to 0.5 ng/g for tri-CB on GCECD. Because ECNI-MS has a higher detectability of highly chlorinated PCB congeners, measurements were performed only for tetra-CB to octa-CB for which the determination limits ranged between 0.01 and 0.2 ng/g wet wt. The determination limits for non-ortho PCBs were 2, 0.4, and 0.4 pg/g wet wt for PCB 77, PCB 126, and PCB 169, respectively. Chiral PCBs were analyzed on an Agilent 6890 GC coupled with a 5973 MSD with EI in SIM mode. The enantiomeric separation was carried out on a 30 m 0.25-mm i.d. Chirasil-Dex column with a 0.25mm film thickness (Chrompack, Raritan, NJ, USA). To reduce the column bleeding, a 3 m 0.25-mm i.d. DBXLB column was connected to the main column with a Press-Tight connector (Agilent, Palo Alto, USA). The injector and detector temperatures were 200 and 2501C, respectively. A 5 5-mL aliquot was injected in large volume injection mode in a manner similar to that described for coplanar PCBs determination. The quadrupole mass spectrometer was operated at an electron energy of 70 eV. Two characteristic ions of each PCB homologue were monitored. The enantiomeric ratio was defined in this study as the proportion of the peak area of the first to the second eluting atropisomer peak (E1/E2), regardless of whether its optical rotation was known or not. For the validation of the chiral analytical procedure, enatiomeric ratios (mean7SD) of a-HCH and four PCB congeners were determined on MSD by five successive injections of a standard solution at concentration of 100 pg/mL: aHCH (0.9970.02), PCB 84 (1.0170.01), PCB 95 (1.0070.01), PCB 149 (1.070.1), and PCB 132 (1.0170.02).
3. Results and discussion GC-ECD and GC/MSD are both widely used for PCB and pesticide determination in environmental samples. While ECD is more suitable for routine environmental investigation, ECNI-MS can give more structural information and can detect lower amounts of highly chlorinated congeners, but fails to have a high detectability for lower chlorinated congeners. However, in the analyzed human tissues, the concentrations of low-chlorinated congeners (o3 chlorine atoms) were very low. Because data obtained on the two detectors matched very well (o15% difference in most cases), the concentrations of pollutants in human tissues were given based on the determination by ECNI-MS. Table 2 summarizes the PCB congener specific and organochlorine pesticide concentrations (expressed on a wet weight basis) in tissues obtained from the 11 subjects. The concentration of extractable lipids in the samples was low, ranging from 0.44% (brain) to 13.8% (liver).
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Table 2 Summary data for PCBs and organochlorine pesticides in human tissues (ng/g wet wt) LODa
Lipids (%)
Liver (n ¼ 11)
Muscle (n ¼ 3)
Kidney (n ¼ 3)
Mean
Range
Mean
Range
Mean
Range
5.3
2.6–13.8
2.7
1.4–3.7
2.7
1.8–3.6
0.4
0.90 ND 0.12 0.82 ND ND 12.2 ND ND ND ND
ND–1.1
0.70 ND 0.26 1.7 ND ND 4.8 ND ND 0.24 ND
ND–1.5
1.3 ND ND ND ND ND ND ND ND ND ND
HCB a-HCH g-HCH b-HCH d-HCH o; p0 -DDE p; p0 -DDE o; p0 -DDD o; p0 -DDT p; p0 -DDD p; p0 -DDT
0.35 0.10 0.10 0.10 0.10 0.20 0.20 0.20 0.20 0.20 0.20
2.2 0.05 0.20 2.3 ND ND 20.9 ND ND 0.83 0.19
0.52–5.7 ND–0.53 ND–0.41 ND–13.2
PCB74 PCB95 PCB101 PCB99 PCB110 PCB149 PCB118 PCB132 PCB153 PCB105 PCB163 PCB138 PCB187 PCB183 PCB128 PCB177 PCB167 PCB156 PCB180 PCB189 PCB199 PCB170 PCB196 PCB194
0.10 0.10 0.10 0.10 0.10 0.10 0.10 0.10 0.20 0.05 0.05 0.20 0.20 0.10 0.05 0.05 0.05 0.05 0.20 0.01 0.10 0.20 0.10 0.10
0.24 ND 0.18 0.58 ND ND 0.73 0.54 7.5 0.16 1.3 3.9 1.4 0.71 0.06 0.59 0.18 0.84 5.73 0.02 0.50 2.5 0.82 0.91
ND–0.86
29.4 579
P PCBs (ng/g wet wt) P PCBs (ng/g lipid wt)
3.5–36.9
0.74–2.9 ND–2.1
ND–0.37 ND–2.5
1.7–19.0
ND–0.50
ND–0.34 ND–1.9 ND ND–0.15 0.21–1.8 0.54–1.6 0.69–23.2 ND–0.42 0.19–3.1 0.46–11.3 ND–3.3 0.11–2.0 ND–0.25 0.14–1.1 ND–0.33 0.08–2.5 0.42–18.5 ND–0.05 ND–1.4 0.22–7.4 ND–2.6 ND–2.8
0.17 ND 0.08 0.33 ND ND 0.53 0.26 9.5 0.11 1.5 3.9 1.6 0.69 ND 0.36 0.17 1.2 8.6 0.06 0.68 3.1 1.1 1.4
2.7–84.8 70–1462
35.3 1139
ND–0.39 ND–4.1
1.3–8.5
ND–0.72
Brain (n ¼ 1)
0.09–0.49 0.04–0.29 0.21–2.0 1.8–13.4 ND–0.12 0.17–1.1 0.62–4.7 0.25–1.6 0.41–2.3
ND ND 0.10 ND ND ND 0.22 0.47 2.6 0.07 0.47 1.1 0.51 0.12 ND 0.13 0.07 0.36 2.4 0.02 0.21 0.87 0.30 0.42
ND–0.49 ND–0.78 0.62–6.3 ND–0.10 0.13–1.1 0.46–2.4 ND–1.3 ND–0.36 ND–0.06 0.07–0.25 0.04–0.15 0.13–0.93 0.37–6.0 ND–0.04 ND–0.52 ND–2.3 ND–0.76 ND–1.0
ND ND ND ND ND ND 0.34 0.36 3.10 0.09 0.51 1.36 0.61 0.19 ND ND 0.07 0.38 2.78 ND 0.27 1.00 0.36 0.42
7.1–50.8 502–1539
10.6 386
3.2–24.5 90–840
11.8 2691
ND–0.24 ND–0.99
0.10–0.93 ND–0.79 1.7–13.6 ND–0.19 0.29–2.4 0.69–6.4 0.43–2.6 0.14–1.1
ND–0.29
ND, not determined. a LOD, limit of determination.
3.1. Organochlorine pesticides and PCBs in human tissues The most abundant pesticide residue in the studied human tissue was p; p0 -DDE, and it could be detected in all human tissues, while p; p0 -DDT could only be found in one liver sample. As liver is the principal organ for metabolization, p; p0 -DDD could be determined in 7 of 11 liver samples, while its concentration was below the limit of determination in all other samples save one kidney sample. Biochemically, p; p0 -DDT is
dechlorinated to p; p0 -DDD, which can then be either metabolized further to p; p0 -DDE or directly excreted from the body (Kutz et al., 1976). However, it may be currently assumed that the presence of p; p0 -DDE in human tissues is not derived from direct ingestion of p; p0 -DDT, but rather from the ingestion of p; p0 -DDE previously degraded in the environment (DuarteDavidson et al., 1994). In our investigation, the concentrations of p; p0 -DDE and HCB in the liver samples from Belgium ranged from 3.5 to 36.9 ng/g wet wt (or 93–1417 ng/g lipid) and from
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0.52 to 5.7 ng/g wet wt (or 14–156 ng/g lipid), respectively (Table 2). These data were similar to those found in Sweden (Guvenius et al., 2001), with concentrations of p; p0 -DDE and HCB in human liver samples ranging from 155 to 1662 and from 17 to 156 ng/g lipid, respectively. b-HCH and g-HCH were found in all liver samples and in some muscle and kidney samples. The high ratio of b-HCHXSHCH indicated that b-HCH is the most persistent and metabolically stable HCH isomer and that the investigated specimens had suffered technical HCH pollution in the past. Although a-HCH is predominant in technical HCH, the concentration of a-HCH was much lower than that of b-HCH due to faster degradation in the body. The relatively high concentration of g-HCH in some subjects revealed that the wide use of lindane is currently one of the main sources of HCH pollution. Except for adipose tissue, there are very few data available regarding the investigation of PCB in human tissues. Table 3 lists some data reported on the concentration of PCBs in human tissues from European populations. Corrigan et al. (1996) investigated PCB residues in brain samples obtained from two men with Parkinson’s disease and from controls. PCBs were not found in the frontal cortex of patients or controls. In a recent study, the levels of PCBs in adipose tissue and liver from one woman and four men in Sweden ranged from 475 to 2085 ng/g lipid (Guvenius et al., 2001). These concentrations are comparable with the concentrations found in human liver samples from Belgium (Tables 2 and 3). To estimate human PCB and pesticide body burden, adipose tissues were widely investigated (Alawi et al., 1999; Corsolini et al., 1995; Falandysz et al., 1994; Focardi et al., 1986; Greve and Van Zoonen, 1990; Kang et al., 1997; Waliszewski et al., 1998). The mean concentrations of DDE and PCBs in Finnish adipose tissue were 567 and 504 ng/g lipid, respectively (Smeds
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and Saukko, 2001). In a general population from Wales (UK), SPCB and SDDT concentrations in human adipose tissues varied between 0.2 and 1.8 mg/g and between 0.11 and 5.6 mg/g tissue, respectively (DuarteDavidson et al., 1994). Twenty human adipose tissue samples from Belgium were analyzed in 2000 for PCBs and pesticides (Covaci et al., 2002), with the median value for the sum of the PCBs (35 congeners) being 841 ng/g lipid wt and ranging from 286 to 1802 ng/g lipid wt. The median concentration of HCB was 43 ng/g lipid wt and ranged from 13 to 85 ng/g lipid wt. The median of the sum of the DDTs was 290 and ranged from 47 to 2802 ng/g lipid wt (Covaci et al., 2002). These values were much lower than those in American populations (Mariottini et al., 2000; Robinson et al., 1990) but in the same range as those found in European populations (Guvenius et al., 2001; Smeds and Saukko, 2001). Direct comparisons of data obtained from liver samples on a wet weight basis with those obtained from adipose tissues is meaningless. It was reported that SPCB levels in liver were 0.8 times high than SPCB levels in adipose tissues if the concentrations were expressed on a lipid basis (Lanting et al., 1998). Due to their low lipid content, it had been considered that muscle, kidney, and brain tissues make little contribution to the human body burden (Lanting et al., 1998). However, muscle is the main body tissue; therefore, investigation of the pollutant residues in this tissue was deemed necessary. Unexpectedly, the concentrations of PCBs and pesticides in human muscle were in the same range as those found in liver. For some animals, the concentrations of PCBs and DDTs in liver were a few times or one order of magnitude higher than those found in muscle (Al-Mohanna, 2000; Metcalfe et al., 1999; Zitko et al., 1998). This phenomenon was also reported by Bachour et al. (1998). They found that the concentration of highly chlorinated congeners was significantly higher on a lipid basis in muscle than in any of the other organs examined in humans. This
Table 3 Previously reported data for PCBs in human adipose tissue and liver (ng/g lipid) Country
Location
Year
Finland
Helsinki
1983 1982–1983
Finland Poland UK Italy Italy Sweden Belgium Belgium a
Turku Skierniewice Gdansk Wales Siena Siena
Antwerp Antwerp
On wet basis concentration.
1979 1990 1990–1991 1991–1992 1996–1997 1994 2000 2000–2001
Age (years) 15–91 19–95 27–91 35–68
21–83
19–77 5–76
Tissue
Range
Mean
Reference
Adipose Liver Adipose Adipose Adipose Adipose Adipose Adipose Adipose Liver Adipose Liver
80–1100 ND–9100 9.6–1996 750–1900 760–4700 257–2628
380 1100 504
Mussalo-Rauhamaa (1991)
931 1720 493a
Duarte-Davidson et al. (1994) Corsolini et al. (1995) Mariottini et al. (2000) Guvenius et al. (2001)
841 579
Covaci et al. (2002) Present study
364–686a 561–2343 475–2085 286–1802 70–1462
Smeds and Saukko (2001) Falandysz et al. (1994)
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situation can be explained as follows: for some predators, such as whales or seals, the daily intake of PCBs and DDTs is much higher than their elimination ability. In this case, liver, as a main organ for metabolism with a high blood flow rate, represents a depot organ for PCBs and DDTs. Bioaccumulation in terrestrial trophic chains was reported to be less accentuated than in the marine environment (Kumar et al., 2001). Being an omnivorous mammal, humans are exposed to PCBs primarily via low-level food contamination. Even in early investigations, market basket studies in the USA have shown a background PCB daily intake of 14 ng/kg/day (Gartrell et al., 1986). In the last decade, PCB contamination has decreased significantly, and data on PCB residues in human adipose tissues suggest that a steady state has been reached (Smeds and Saukko, 2001). Differing from liver, muscle tissue represents a comparatively inactive compartment with a low rate of metabolism. When the intake and elimination in humans has reached a steady state, the distribution of lipophilic compounds reaches an equilibrium and depends mostly on the transport mechanism. The concentrations of PCBs and DDTs in kidney and brain were significantly lower than in liver and muscle. Bachour et al. (1998) investigated 25 deceased males and females and found that the concentration of PCBs in brain was lower than in liver and muscle, especially for highly chlorinated congeners. Although a slightly lower concentration of SPCB was observed in the livers of children (samples 6L and 7L), the number of samples analyzed was too small to examine age- and genderspecific differences in PCB concentrations. While some data suggested an increase in concentrations with age (Robinson et al., 1990; Smeds and Saukko, 2001), other studies showed no such differences in PCB
serum
adipose
liver
concentrations with age (Kang et al., 1997). It seems that the relationship between the SPCB levels in human tissue is not as simple as we expected (Mariottini et al., 2000). In our study, subjects 4–6 were from the same family and therefore dietary differences could be neglected. While the ages of individual members were 5, 31, and 35 years, the total PCB concentrations in liver were similar for all three specimens (11.2, 14.4, and 12.7 ng/g wet wt, respectively). 3.2. Distribution of PCB congeners in human tissues The PCB congener patterns in different organs from individual subjects were similar (Table 2). The PCB pattern in our subjects and the previously reported distribution of PCB congeners in adipose tissue (Covaci et al., 2002) and serum (Covaci, 2002) are shown in Fig. 1. There were no significant differences between the distribution of congeners with different degrees of chlorination. The most abundant PCB congeners were PCB 153 and PCB 180, each of them accounting for approximately 30% and 25%, respectively, of the sum PCB in all samples. It is well know that the PCB pattern shifts from low-chlorinated congeners to higher chlorinated congeners when the organisms move from lower to higher trophic levels (Jones, 1988; Kumar et al., 2001). The higher concentrations of PCB 153 and PCB 180 (persistent congeners) and the absence of lowchlorinated congeners in individual subjects indicated that the principal source of contamination with PCBs was from diet and not from direct exposure. It is most likely that the contaminants passed several bioaccumulation steps before being ingested. Lanting et al. (1998) analyzed PCBs in adipose tissue, liver, and brain of nine fetuses that died in uteri. The
muscle
kidney
brain
40
% of sum 9 PCB congeners
35 30 25 20 15 10 5 0 PCB105
PCB118
PCB138
PCB153
PCB156
PCB170
PCB180
PCB187
PCB194
Fig. 1. Distribution of selected PCB congeners in human serum and different tissues (muscle, kidney, liver, and brain).
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medians (ranges) for the sum of PCBs 118, 138, 153, and 180 in the tissues were: adipose tissue 235 (97–768), liver 198 (67–362), and brain 50 (22–122) ng/g lipid. There were no differences in the distributions of congeners 118, 138, 153, and 180 between the three tissues and there was no statistically significant association between tissue PCB levels and gestational age (Lanting et al., 1998). We noticed that congeners with less than three chlorine atoms were not found in the samples. This phenomenon is different from those found in previous investigations. In 1991, the National Human Adipose Tissue Survey revealed that infants and children (age o14) had detectable levels of tri-CB but no nona-CBs. However, in the group aged 15–44 tri-CBs were not detectable but nona-CBs were (Phillips and Birchard, 1991). 3.3. Non-ortho PCB congeners Although non-ortho PCBs accounted for only 0.1% of SPCBs, these congeners are of great concern because of their dioxin-like toxicity. Similar to 2,3,7,8-TCDD, they bind to the cytosolic aryl hydrocarbon receptor to build a substrate–receptor complex that can enter the cell nucleus and interfere with gene expression. Thus, monitoring of toxic non-ortho PCBs is necessary for understanding their possible impact on human health (Corsolini et al., 1995). Data for non-ortho PCBs in studied human tissues are presented in Table 4. Normally, a decrease in the concentrations of the three non-ortho PCBs with an increasing degree of chlorination was found in water, soil, and sediment samples. PCB 77 was also abundant
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in some technical mixtures (Erickson, 1997) and in animals which were situated at the bottom of the food chain. PCB 126 is abundant in some predators, such as birds, fox, and marine mammals, while PCB 169 is occasionally found in fish, birds, and some terrestrial animals, with a lower concentration relative to the other two congeners (Kumar et al., 2001). However, in our subjects this distribution was different from those found in other terrestrial mammals. In 10 of 18 tissue samples, the concentrations of PCB 169 were higher than those of PCB 126 (Table 4). This distribution of non-ortho PCBs in human tissue may have a time trend. Choi et al. (2002) investigated non-ortho PCB concentrations in Japanese human adipose tissue. The concentrations (mean7SD) of PCB 126 and PCB 169 in 1970–1971 were 2027146 and 50725 pg/g lipid wt, respectively, while in 2000 they were 72760 and 61725 pg/g lipid wt. Patterson et al. (1994) determined non-ortho PCBs in human adipose tissues from Atlanta. The mean concentrations of PCB 126 and PCB 169 were 102 and 67 pg/g lipid, respectively. This distribution might have been due to the fact that previously high PCB levels might have induced detoxifying hepatic enzymes (cytochrome P450) that lead to the degradation of less persistent congeners, such as PCB 77 and PCB 126 (Corsolini et al., 1995; Kannan et al., 1993; Patterson et al., 1994). The relative abundance of the three non-ortho PCB congeners in different tissues is similar in liver, muscle, and kidney. The mean concentration of PCB 77, PCB 126, and PCB 169 were 31.0, 56.6, and 59.1 pg/g lipid wt in liver samples. These values are lower than those
Table 4 Concentration of non-ortho PCBs in human tissues (pg/g wet weight) Subject no. 1
2
3
4 5 6 7 8 9 10 11 a b
Sample no.
Lipids (%)
PCB 77 (TEF:0.0001)
PCB 126 (TEF:0.1)
PCB 169 (TEF:0.01)
No-TEQa (pg/g wet wt)
Mo-TEQb (pg/g wet wt)
1B 1K 1L 1M 2K 2L 2M 3K 3L 3M 4L 5L 6L 7L 8L 9L 10L 11L
0.4 2.9 3.6 3.7 1.8 5.8 1.4 3.6 5.8 3.1 3.6 4.4 5.5 3.8 2.6 13.8 3.7 5.7
2.6 2.2 o2.0 2.2 o2.0 o2.0 o2.0 2.8 o2.0 o2.0 o2.0 o2.0 o2.0 o2.0 o2.0 2.4 o2.0 5.8
0.6 1.3 3.5 2.8 o0.4 1.1 o0.4 2.2 1.9 1.1 2.5 2.6 3.5 1.4 2.2 3.9 3.7 1.7
0.4 3.1 4.2 5.1 o0.4 2.1 0.5 3.2 6.3 3.3 1.3 0.8 0.8 o0.4 3.8 4.7 2.1 1.2
0.07 0.16 0.39 0.33 o0.04 0.13 o0.04 0.25 0.25 0.14 0.27 0.27 0.36 0.14 0.26 0.43 0.39 0.18
0.23 0.52 0.93 0.54 0.07 0.38 0.12 0.03 1.37 0.77 0.19 0.27 0.24 0.07 0.51 1.08 0.54 0.41
The No-TEQ values were calculated from the concentrations of non-ortho PCB congeners. The Mo-TEQ values were calculated from the concentrations of mono-ortho PCB congeners 105, 118, 156, and 189.
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previously reported in adipose tissue from Italy, Japan, Poland, and North America (Corsolini et al., 1995; Greve and Van Zoonen, 1990; Tanabe et al., 1989; Williams and Lebel, 1991). Toxic equivalent (TEQ) values based on the most recent toxic equivalency factors (TEFs) (van den Berg et al., 1998) were calculated to estimate the relative contributions of toxic potential from congeners know to cause induction in the detoxifying enzyme activities of the P450 family. They ranged from o0.04 to 0.43 and from 0.07 to 1.37 pg/g TEQ wet wt for non-ortho and mono-ortho PCBs, respectively (Table 4). The mean TEQ values were 0.28, 0.22, 0.21, and 0.07 pg/g wet wt in liver, muscle, kidney, and brain, respectively. The TEQ of the non-ortho PCBs was lower than that of mono-ortho congeners. TEQ values calculated for the three main mono-ortho PCBs (PCB 105, PCB 118, and PCB 156) were 0.51, 0.66, 0.31, and 0.23 pg/g wet wt in liver, muscle, kidney, and brain, respectively. The highest contribution was from PCB 156, which accounted for about 60% of the total PCB TEQ. 3.4. Enantiomeric composition of chiral PCBs and aHCH in human tissue The ERs for chiral PCBs and a-HCH investigated in this study are given in Table 5. The concentration of PCB 84 was very low in all samples. Due to the low signal-to-noise ratio and poor resolution, reliable quantification for PCB 84 atropisomer in the samples was impossible. In 3 of 18 samples (5L, 10L, and 11L), the enantiomeric ratio of a-HCH could be determined (ER=1.06, 1.09, and 1.02, respectively). Racemic or nearly racemic ratios for chiral PCBs were also found in
all muscle, kidney, and brain samples, ranging from 1.02 to 1.32 for PCB 95, 1.00 to 1.18 for PCB 149, and 0.88 to 1.09 for PCB 132. Although some papers suggested that enantioselective enrichment might occur in the brain (Vetter and Schurig, 1997), there was no enrichment of chiral PCBs in the brain sample. Most of the enantiomeric enrichment was found in liver samples. The mean (range) ERs in liver were 1.69 (1.04–2.97), 1.16 (0.99– 1.41), and 0.74 (0.48–0.97) for PCB 95, PCB 149, and PCB 132, respectively. The first attempt to determine the enantiomeric ratio of chiral PCB in human fluids was done by Glausch et al. (1994). Subsequent research has indicated that the second-eluted enantiomer of PCB 132 was enriched in human milk extracts (Glausch et al., 1995). Since the elution order of PCB 132 atropisomers on Chirasil-Dex was known (Haglund and Wiberg, 1996), (+)-PCB132 was the most abundant enantiomer in human milk. Because the same column (Chirasil-Dex) was used for chiral separation in our work, the results indicate that, similarly to milk, an enrichment of (+)-PCB 132 atropisomers was observed in human liver. All ERs of PCB 95 and PCB 149 were 41 or approximately equal to 1, while most of ERs for PCB 132 were o1 or close to 1. The elution order for PCB 95 and PCB 149 atropisomers on the Chirasil-Dex column might be reversed with that of PCB 132 atropisomers, but these results should be confirmed with pure PCB 95 and PCB 149 atropisomers. Relationships between the ERs of three chiral PCBs are shown in Figs. 2a and b and suggest that the bioselective metabolism of chiral PCBs has the same trend, although the ratio is different. The ERs for individual chiral PCBs in samples from 4L, 5L, 3.50
Subject no.
Sample no.
ER (a-HCH)
ER (PCB 95)
ER (PCB 149)
ER (PCB 132)
1.02 1.02 1.04 1.03 1.03 1.26 1.06 1.32 1.17 1.12 1.84 2.95 2.97 1.29 1.69 2.10 1.14 1.18
1.05 1.01 0.99 1.00 1.02 1.04 1.03 1.18 1.02 1.02 1.32 1.41 1.39 1.06 1.29 1.22 1.00 1.04
1.09 1.06 0.97 0.99 1.00 0.81 1.02 0.95 0.93 0.88 0.66 0.50 0.48 0.97 0.65 0.50 0.91 0.78
3.00
ER(PCB 95)
Table 5 Enantiomeric ratios of chiral PCBs and a-HCH in human tissues
2.50
r = 0.9433, p<0.001
2.00 1.50 1.00
2
3
4 5 6 7 8 9 10 11
1B 1K 1L 1M 2K 2L 2M 3K 3L 3M 4L 5L 6L 7L 8L 9L 10L 11L
1.06
1.09 1.02
0.50 0.50
1.00
1.50
2.00
2.50
1/ER(PCB 132)
(a) 1.50 1.40
ER(PCB 149)
1
1.30
r = 0.8674, p<0.001
1.20 1.10 1.00 0.90 0.80
(b)
1.00
1.20
1.40
1.60
1.80
2.00
2.20
1/ER(PCB 132)
Fig. 2. Relationships between ER of PCB 95 with 1/ER of PCB 132 (a) and ER of PCB 149 with of 1/ER 132 (b) in human tissues.
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and 6L (Table 1) were similar and showed that they were more depended on daily intake and not on age or sex.
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Adipose Tissue survey. Arch. Environ. Contam. Toxicol. 21, 159–168. Ramos, L., Hernandez, L.M., Gonzalez, M., 1999. Simultaneous separation of coplanar and chiral polychlorinated biphenyls by offline pyrenyl-silica liquid chromatography and gas chromatography, enantiomeric ratios of chiral congeners. Anal. Chem. 71, 70–77. Robinson, P.E., Mack, G.A., Remmers, J., Levy, R., Mohadjer, L., 1990. Trends of PCB, hexachlorobenzene, and b-benzene hexachloride levels in the adipose tissue of the US population. Environ. Res. 53, 175–192. Schecter, A., 1998. Selective historical review of congener-specific human tissue measurements as sensitive and specific biomarkers of exposure to dioxins and related compounds. Environ. Health Perspect. 106, 737–742. Smeds, A., Saukko, P., 2001. Identification and quantification of polychlorinated biphenyls and some endocrine disrupting pesticides in human adipose tissues from Finland. Chemosphere 44, 1463–1471. Tanabe, S., Kannan, N., Wakimoto, T., Tatsukawa, R., Okamoto, T., Masuda, Y., 1989. Isomer specific determination and toxic evaluation of potentially hazardous coplanar PCBs, dibenzofurans and dioxins in the tissue of ‘‘Yusho’’ PCB poisoning victim and in the causal oil. Toxicol. Environ. Chem. 24, 215–231. van den Berg, M., Birnbaum, L.S., Bosveld, A.T.C., Brunstrom, B., Cook, P., Feeley, M., Giesy, J.P., Hanberg, A., Hasegawa, R., Kennedy, S.W., Kubiak, T., Larsen, J.C., van Leeuwen, F.X.R., Liem, A.K.D., Nolt, C., Peterson, R.E., Poellinger, L., Safe, S.,
Schrenk, D., Tillitt, D., Tysklind, M., Younes, M., Waern, F., Zacharewksi, T., 1998. Toxic equivalency factors (TEFs) for PCBs, PCDDs, and PCDFs for humans and wildlife. Environ. Health Perspect. 106, 775–792. van Larebeke, N., Hens, L., Schepens, P., Covaci, A., Baeyens, J., Everaert, K., Bernheim, J.L., Vlietinck, R., De Poorter, G., 2001. The Belgian PCB and dioxin incident of January–June 1999: exposure data and potential impact on health. Environ. Health Perspect. 109, 265–273. Vetter, W., Schurig, V., 1997. Enantioselective determination of chiral organochlorine compounds in biota by gas chromatography on modified cyclodextrins. J. Chromatogr. A 774, 143–175. Waliszewski, S.M., Aguirre, A.A., Infanzon, R.M., Rivera, J., Infanzon, R., 1998. Time trend of organochlorine pesticide residues in human adipose tissue in Veracruz, Mexico: 1988–1997 survey. Sci. Total Environ. 221, 201–204. Williams, D.T., Lebel, G.L., 1991. Coplanar polychlorinated biphenyl residues in human adipose tissue samples from Ontario municipalities. Chemosphere 22, 1019–1028. Wong, C.S., Garrison, A.W., Smith, P.D., Foreman, W.T., 2001. Enantiomeric composition of chiral polychlorinated biphenyl atropisomers in aquatic and riparian biota. Environ. Sci. Technol. 35, 2448–2454. Zitko, V., Stenson, G., Hellou, J., 1998. Levels of organochlorine and polycyclic aromatic compounds in harp seal beaters (Phoca groenlandica). Sci. Total Environ. 221, 11–29.
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Levels and chiral signatures of persistent organochlorine pollutants in human tissues from Belgium Shaogang Chu, Adrian Covaci, and Paul Schepens Toxicological Center, University of Antwerp, Universiteitsplein 1, Antwerpen (Wilrijk) 2610, Belgium Received 8 August 2002; received in revised form 7 January 2003; accepted 17 January 2003
Abstract Polychlorinated biphenyls (PCBs) and organochlorine pesticides (OCPs) were measured in human tissue samples including muscle, liver, brain, and kidney. The samples were obtained at autopsy in 2000–2001 from three women and eight men from Belgium, aged between 5 and 76 years. The measured PCBs included 23 ortho-substituted congeners and 3 non-ortho-substituted congeners (PCB 77, PCB 126, and PCB 169). The mean concentrations of SPCBs were 29.4, 35.3, 10.6, and 11.8 ng/g wet wt in liver, muscle, kidney, and brain, respectively. HCB, g-HCH, b-HCH, p; p0 -DDE, and p; p0 -DDD were found in all samples, while p; p0 DDT could only be found in one liver sample. The most abundant pesticide was p; p0 -DDE. PCB 153 and PCB 180 were the main ortho-substituted congeners found in all the samples, while the concentration of the congeners with less than three chlorine atoms was below the limit of determination. In 10 of 18 tissues, the concentrations of PCB 169 were higher than the concentration of PCB 126. These results are consistent with the order of half-life of these congeners in humans and indicate that a steady state had been reached in these subjects. The enantiomeric compositions of a-HCH and chiral PCBs, including PCB 95, PCB 132, and PCB 149, were also measured. a-HCH was found to be racemic in three liver samples, while chiral PCB 95, PCB 149, and 132 showed racemic or nearly racemic compositions in muscle, kidney, and brain. Higher enatiomeric ratios (ERs) for the three chiral PCBs were found in liver samples. The mean (range) ERs in liver were 1.69 (1.04–2.97), 1.16 (0.99–1.41), and 0.74 (0.48–0.97) for PCB 95, PCB 149, and PCB 132, respectively. r 2003 Elsevier Science (USA). All rights reserved.
1. Introduction Polychlorinated biphenyls (PCBs) and organochlorine pesticides (OCPs) are ubiquitous pollutants that can be biologically amplified in food webs to high concentrations. Despite restrictions on their manufacture and use in most industrialized countries since the 1970s, they are still of public concern and accidents may occasionally occur (Chen et al., 1980; Kuratsune et al., 1972). In January 1999, 500 tons of feed contaminated with approximately 50 kg of PCBs and 1 g of dioxins were distributed to animal farms in Belgium (van Larebeke et al., 2001). An extensive investigation of possible effects caused by the incident has been done in the last 2 years (van Larebeke et al., 2001). Significant advancements have been introduced in PCB analysis, which directly impacted the validity of reported trend analyses. In the last decade, more
Corresponding author. Fax: +32-38202722. E-mail address: [email protected] (A. Covaci).
attention has been paid to the determination of nonortho PCBs, which show similar toxicity to polychlorinated dibenzo-p-dioxins and dibenzofurans (Schecter, 1998). More recently, attention has been focused on PCBs and other organochlorine contaminants that display axial chirality, and it was suggested that the enantiomeric ratio of chiral contaminants in different species may give additional information on possible bioaccumulation and degradation pathways (Kallenborn and Huhnerfuss, 2001; Wong et al., 2001). While there is now evidence that PCB concentrations in the environment were generally declining last decade, data available on humans suggest that the declining was much slower to respond to the restriction in use or incident. Presumably, this is due to the high persistence of PCBs in the fatty tissues in which they are deposited (Erickson, 1997). The data on human contamination with PCB and pesticides are focused on blood or milk, the only biological materials conveniently available from healthy subjects. In most cases, blood reflects the event of actual pollution exposure (Phillips et al., 1989;
0013-9351/03/$ - see front matter r 2003 Elsevier Science (USA). All rights reserved. doi:10.1016/S0013-9351(03)00016-1
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Schecter, 1998) and selective metabolism and excretion may alter the profile. Adipose tissue is another choice for the investigation of human body burden (Phillips and Birchard, 1991). However, chemical behavior and biomagnification may vary among organs of the studied biota (Boon et al., 1987). Because human organs are not convenient to obtain, relatively limited data on the levels of PCBs and in particular of non-ortho PCBs in human tissues (other than adipose tissue) are now available. Moreover, data on the enantiomeric composition of PCB atropisomers in human tissues have not been reported at this time. The aim of this study was to investigate the levels and distribution of PCBs (including non-ortho congeners) and OCPs in selected human tissues from the Belgian general population and to explore eventual relationships with the enantiomeric ratios (ERs) of chiral pollutants.
Table 1 Identification of human tissue samples from Belgium
2. Materials and methods 2.1. Samples The sample materials consisted of postmortem human tissues from 11 deceased males and females (age range 5–76 years) that died suddenly from causes unrelated to environmental contaminants. All samples were collected in 2000–2001 at one Belgian hospital. Parts of the brain, liver, kidney, and muscle were obtained from unembalmed cadavers during autopsies and frozen at 201C until analysis. Sample details are listed in Table 1. Subjects 4–6 came from the same family. 2.2. Chemicals and standards All solvents were of pesticide residue analysis from Merck (Wessel, Germany). Individual PCB congeners and pesticide standards were purchased from Dr. Ehrenstorfer (Augsburg, Germany), while 13C12-labeled non-ortho-substituted PCBs (PCB 77, 126, 169) were obtained from Cambridge Isotope Laboratories (Andover, MA, USA). 2.3. Extraction, clean-up and fractionation Eight grams of tissue samples was ground with sufficient Na2SO4 to disintegrate the sample and bind the water. Internal standards (PCB 143, PCB 46, and eHCH for the determination of ortho-substituted PCBs and OCPs; 13C12-labeled PCB 77, 126, and 169 for the determination of non-ortho PCBs) were added to each tissue and blank sample prior to extraction. The samples were extracted for 4 h with hexane/acetone (3:1, v/v) in a BUCHI B-811 Soxhlet extraction system (Bruxelles, Belgium) operated in hot extraction mode. After concentration to approximately 5 mL, an aliquot
Subject no.
Sex
Age
Sample no.
Tissue
Lipid content (%)
1
Male
Adult
1B 1K 1L 1M
Brain Kidney Liver Muscle
0.4 2.9 3.6 3.7
2
Male
36
2K 2L 2M
Kidney Liver Muscle
1.8 5.8 1.4
3
Male Male Male
Adult
3K 3L 3M
Kidney Liver Muscle
3.6 5.8 3.1
4 5 6
Male Female Male
35 31 5
4L 5L 6L
Liver Liver Liver
3.6 4.4 5.5
7 8 9 10 11
Male Female Male Female Male
Child Child Adult 56 76
7L 8L 9L 10L 11L
Liver Liver Liver Liver Liver
3.8 2.6 13.8 3.7 5.7
of the extract was used to determine the extractable lipids gravimetrically. The remaining extract was cleaned on a column packed with 8 g acid silica (40% sulfuric acid, w/w), followed by a column with 10 g Florisil. One fraction (70 mL hexane) was collected and concentrated to 50 mL. The fractionation of PCB congeners according to their chlorine substitution was achieved by the method reported by Ramos et al. (1999) with some modifications. PCBs were separated by a 250 mm 4.6-mm i.d. Cosmosil 5-PYE column (Nacalai Tesque, Osaka, Japan). The HPLC system consisted of a Gilson 307 pump, 112 UV/VIS detector (Middleton, WI, USA), and a Rheodyne 7125 injector (Rohnert Park, CA, USA) equipped with a 20-mL sample loop. The detector was operated at 254 nm. Hexane was used as a mobile phase at a flow rate of 0.5 mL/min and the separated fractions were collected by a HeliRac LKB2212 fraction collector (LKB-Produkter AB, Bromma, Sweden). The first fraction (6–12 min) contained most ortho-substituted congeners and pesticides, while the second fraction (12–20 min) contained non-ortho-substituted congeners (PCB 77, PCB 126, PCB 169) and mono-ortho-substituted congener PCB 189. For the determination of the enantiomeric composition of PCB atropisomers, the same HPLC system was used to separate the individual chiral PCBs from the other congeners. Heart cut technology was used and fractions of 80 mL were collected for each chiral PCB (PCB 84, PCB 95, PCB 132, and PCB 149) according to their elution times on the HPLC chromatogram.
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2.4. Analysis An Agilent 6890 GC with microcell ECD was used to determine PCBs and pesticides. The separation was performed on a 25 m 0.22 mm, 0.25-mm-film-thickness HT-8 capillary column from SGE (Zulte, Belgium). The carrier and make-up gas were helium and argon/ methane (95:5), respectively. The initial oven temperature was 901C, hold for 1 min, and then to 1801C at 151C/min, hold for 1 min, to 2501C at 31C/min, and to 2901C at 151C and hold for 6 min. The injector temperature and detector temperature were 2701C and 3201C, respectively. A 1-mL extract was injected in pulse splitless mode and the purge time was 1 min. The identification was based on the retention time of individual standard congeners and a multilevel calibration curve was used for quantitative determination. For confirmation, all samples were reanalyzed by GC/ MS with an electron capture negative-impact ionization (ECNI) source and selected ion monitoring mode. An Agilent 6890 GC coupled with Agilent 5973N MSD was used for qualitative and quantitative analyses. A 1-mL aliquot of extract was injected into a 10 m 0.1 mm, 0.10-mm-thickness HT-8 (SGE) narrow-bore capillary column in solvent vent injection mode. The injection parameters were: initial temperature 601C, hold for 1 min, and then to 3001C at 6001C/min and hold for 20 min. The vent time was 0.9 min. The initial oven temperature was 601C, hold for 3 min, and then to 1501C at 501C/min, to 2501C at 101C/min, and to 2901C/min at 201C/min and hold for 10 min. The ion source and quadrupole temperatures were 2001C and 1501C, respectively. Methane was used as the reagent gas. Three characteristic ions of each analyte (tetra- to octa-CB and pesticides) were selected with seven retention windows. Identification was based on retention times and the intensity of the acquired ions in sample peaks within 10% of the mean values obtained from the corresponding standards. Quantitative determination was done using five-level calibration curves obtained from a standard solution containing all congeners. The method used for the analysis of non-ortho PCBs was similar to that described above, except that the injector system was operated in large volume injection mode. A 5 5-mL aliquot extract was injected into injector and the purge time was 2.2 min. Procedural blanks were subjected to the entire analytical procedure to determine background interference, and peaks were quantified only when the signal/ noise ratio was 43. The mean recovery of internal standards in individual samples was 75.1%. The limits of determination varied for individual PCB congeners and detector, and these generally decreased with increased levels of chlorination. Limits of determination for ortho-substituted PCBs and pesticides ranged from
169
0.01 ng/g for penta-CB to 0.5 ng/g for tri-CB on GCECD. Because ECNI-MS has a higher detectability of highly chlorinated PCB congeners, measurements were performed only for tetra-CB to octa-CB for which the determination limits ranged between 0.01 and 0.2 ng/g wet wt. The determination limits for non-ortho PCBs were 2, 0.4, and 0.4 pg/g wet wt for PCB 77, PCB 126, and PCB 169, respectively. Chiral PCBs were analyzed on an Agilent 6890 GC coupled with a 5973 MSD with EI in SIM mode. The enantiomeric separation was carried out on a 30 m 0.25-mm i.d. Chirasil-Dex column with a 0.25mm film thickness (Chrompack, Raritan, NJ, USA). To reduce the column bleeding, a 3 m 0.25-mm i.d. DBXLB column was connected to the main column with a Press-Tight connector (Agilent, Palo Alto, USA). The injector and detector temperatures were 200 and 2501C, respectively. A 5 5-mL aliquot was injected in large volume injection mode in a manner similar to that described for coplanar PCBs determination. The quadrupole mass spectrometer was operated at an electron energy of 70 eV. Two characteristic ions of each PCB homologue were monitored. The enantiomeric ratio was defined in this study as the proportion of the peak area of the first to the second eluting atropisomer peak (E1/E2), regardless of whether its optical rotation was known or not. For the validation of the chiral analytical procedure, enatiomeric ratios (mean7SD) of a-HCH and four PCB congeners were determined on MSD by five successive injections of a standard solution at concentration of 100 pg/mL: aHCH (0.9970.02), PCB 84 (1.0170.01), PCB 95 (1.0070.01), PCB 149 (1.070.1), and PCB 132 (1.0170.02).
3. Results and discussion GC-ECD and GC/MSD are both widely used for PCB and pesticide determination in environmental samples. While ECD is more suitable for routine environmental investigation, ECNI-MS can give more structural information and can detect lower amounts of highly chlorinated congeners, but fails to have a high detectability for lower chlorinated congeners. However, in the analyzed human tissues, the concentrations of low-chlorinated congeners (o3 chlorine atoms) were very low. Because data obtained on the two detectors matched very well (o15% difference in most cases), the concentrations of pollutants in human tissues were given based on the determination by ECNI-MS. Table 2 summarizes the PCB congener specific and organochlorine pesticide concentrations (expressed on a wet weight basis) in tissues obtained from the 11 subjects. The concentration of extractable lipids in the samples was low, ranging from 0.44% (brain) to 13.8% (liver).
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Table 2 Summary data for PCBs and organochlorine pesticides in human tissues (ng/g wet wt) LODa
Lipids (%)
Liver (n ¼ 11)
Muscle (n ¼ 3)
Kidney (n ¼ 3)
Mean
Range
Mean
Range
Mean
Range
5.3
2.6–13.8
2.7
1.4–3.7
2.7
1.8–3.6
0.4
0.90 ND 0.12 0.82 ND ND 12.2 ND ND ND ND
ND–1.1
0.70 ND 0.26 1.7 ND ND 4.8 ND ND 0.24 ND
ND–1.5
1.3 ND ND ND ND ND ND ND ND ND ND
HCB a-HCH g-HCH b-HCH d-HCH o; p0 -DDE p; p0 -DDE o; p0 -DDD o; p0 -DDT p; p0 -DDD p; p0 -DDT
0.35 0.10 0.10 0.10 0.10 0.20 0.20 0.20 0.20 0.20 0.20
2.2 0.05 0.20 2.3 ND ND 20.9 ND ND 0.83 0.19
0.52–5.7 ND–0.53 ND–0.41 ND–13.2
PCB74 PCB95 PCB101 PCB99 PCB110 PCB149 PCB118 PCB132 PCB153 PCB105 PCB163 PCB138 PCB187 PCB183 PCB128 PCB177 PCB167 PCB156 PCB180 PCB189 PCB199 PCB170 PCB196 PCB194
0.10 0.10 0.10 0.10 0.10 0.10 0.10 0.10 0.20 0.05 0.05 0.20 0.20 0.10 0.05 0.05 0.05 0.05 0.20 0.01 0.10 0.20 0.10 0.10
0.24 ND 0.18 0.58 ND ND 0.73 0.54 7.5 0.16 1.3 3.9 1.4 0.71 0.06 0.59 0.18 0.84 5.73 0.02 0.50 2.5 0.82 0.91
ND–0.86
29.4 579
P PCBs (ng/g wet wt) P PCBs (ng/g lipid wt)
3.5–36.9
0.74–2.9 ND–2.1
ND–0.37 ND–2.5
1.7–19.0
ND–0.50
ND–0.34 ND–1.9 ND ND–0.15 0.21–1.8 0.54–1.6 0.69–23.2 ND–0.42 0.19–3.1 0.46–11.3 ND–3.3 0.11–2.0 ND–0.25 0.14–1.1 ND–0.33 0.08–2.5 0.42–18.5 ND–0.05 ND–1.4 0.22–7.4 ND–2.6 ND–2.8
0.17 ND 0.08 0.33 ND ND 0.53 0.26 9.5 0.11 1.5 3.9 1.6 0.69 ND 0.36 0.17 1.2 8.6 0.06 0.68 3.1 1.1 1.4
2.7–84.8 70–1462
35.3 1139
ND–0.39 ND–4.1
1.3–8.5
ND–0.72
Brain (n ¼ 1)
0.09–0.49 0.04–0.29 0.21–2.0 1.8–13.4 ND–0.12 0.17–1.1 0.62–4.7 0.25–1.6 0.41–2.3
ND ND 0.10 ND ND ND 0.22 0.47 2.6 0.07 0.47 1.1 0.51 0.12 ND 0.13 0.07 0.36 2.4 0.02 0.21 0.87 0.30 0.42
ND–0.49 ND–0.78 0.62–6.3 ND–0.10 0.13–1.1 0.46–2.4 ND–1.3 ND–0.36 ND–0.06 0.07–0.25 0.04–0.15 0.13–0.93 0.37–6.0 ND–0.04 ND–0.52 ND–2.3 ND–0.76 ND–1.0
ND ND ND ND ND ND 0.34 0.36 3.10 0.09 0.51 1.36 0.61 0.19 ND ND 0.07 0.38 2.78 ND 0.27 1.00 0.36 0.42
7.1–50.8 502–1539
10.6 386
3.2–24.5 90–840
11.8 2691
ND–0.24 ND–0.99
0.10–0.93 ND–0.79 1.7–13.6 ND–0.19 0.29–2.4 0.69–6.4 0.43–2.6 0.14–1.1
ND–0.29
ND, not determined. a LOD, limit of determination.
3.1. Organochlorine pesticides and PCBs in human tissues The most abundant pesticide residue in the studied human tissue was p; p0 -DDE, and it could be detected in all human tissues, while p; p0 -DDT could only be found in one liver sample. As liver is the principal organ for metabolization, p; p0 -DDD could be determined in 7 of 11 liver samples, while its concentration was below the limit of determination in all other samples save one kidney sample. Biochemically, p; p0 -DDT is
dechlorinated to p; p0 -DDD, which can then be either metabolized further to p; p0 -DDE or directly excreted from the body (Kutz et al., 1976). However, it may be currently assumed that the presence of p; p0 -DDE in human tissues is not derived from direct ingestion of p; p0 -DDT, but rather from the ingestion of p; p0 -DDE previously degraded in the environment (DuarteDavidson et al., 1994). In our investigation, the concentrations of p; p0 -DDE and HCB in the liver samples from Belgium ranged from 3.5 to 36.9 ng/g wet wt (or 93–1417 ng/g lipid) and from
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0.52 to 5.7 ng/g wet wt (or 14–156 ng/g lipid), respectively (Table 2). These data were similar to those found in Sweden (Guvenius et al., 2001), with concentrations of p; p0 -DDE and HCB in human liver samples ranging from 155 to 1662 and from 17 to 156 ng/g lipid, respectively. b-HCH and g-HCH were found in all liver samples and in some muscle and kidney samples. The high ratio of b-HCHXSHCH indicated that b-HCH is the most persistent and metabolically stable HCH isomer and that the investigated specimens had suffered technical HCH pollution in the past. Although a-HCH is predominant in technical HCH, the concentration of a-HCH was much lower than that of b-HCH due to faster degradation in the body. The relatively high concentration of g-HCH in some subjects revealed that the wide use of lindane is currently one of the main sources of HCH pollution. Except for adipose tissue, there are very few data available regarding the investigation of PCB in human tissues. Table 3 lists some data reported on the concentration of PCBs in human tissues from European populations. Corrigan et al. (1996) investigated PCB residues in brain samples obtained from two men with Parkinson’s disease and from controls. PCBs were not found in the frontal cortex of patients or controls. In a recent study, the levels of PCBs in adipose tissue and liver from one woman and four men in Sweden ranged from 475 to 2085 ng/g lipid (Guvenius et al., 2001). These concentrations are comparable with the concentrations found in human liver samples from Belgium (Tables 2 and 3). To estimate human PCB and pesticide body burden, adipose tissues were widely investigated (Alawi et al., 1999; Corsolini et al., 1995; Falandysz et al., 1994; Focardi et al., 1986; Greve and Van Zoonen, 1990; Kang et al., 1997; Waliszewski et al., 1998). The mean concentrations of DDE and PCBs in Finnish adipose tissue were 567 and 504 ng/g lipid, respectively (Smeds
171
and Saukko, 2001). In a general population from Wales (UK), SPCB and SDDT concentrations in human adipose tissues varied between 0.2 and 1.8 mg/g and between 0.11 and 5.6 mg/g tissue, respectively (DuarteDavidson et al., 1994). Twenty human adipose tissue samples from Belgium were analyzed in 2000 for PCBs and pesticides (Covaci et al., 2002), with the median value for the sum of the PCBs (35 congeners) being 841 ng/g lipid wt and ranging from 286 to 1802 ng/g lipid wt. The median concentration of HCB was 43 ng/g lipid wt and ranged from 13 to 85 ng/g lipid wt. The median of the sum of the DDTs was 290 and ranged from 47 to 2802 ng/g lipid wt (Covaci et al., 2002). These values were much lower than those in American populations (Mariottini et al., 2000; Robinson et al., 1990) but in the same range as those found in European populations (Guvenius et al., 2001; Smeds and Saukko, 2001). Direct comparisons of data obtained from liver samples on a wet weight basis with those obtained from adipose tissues is meaningless. It was reported that SPCB levels in liver were 0.8 times high than SPCB levels in adipose tissues if the concentrations were expressed on a lipid basis (Lanting et al., 1998). Due to their low lipid content, it had been considered that muscle, kidney, and brain tissues make little contribution to the human body burden (Lanting et al., 1998). However, muscle is the main body tissue; therefore, investigation of the pollutant residues in this tissue was deemed necessary. Unexpectedly, the concentrations of PCBs and pesticides in human muscle were in the same range as those found in liver. For some animals, the concentrations of PCBs and DDTs in liver were a few times or one order of magnitude higher than those found in muscle (Al-Mohanna, 2000; Metcalfe et al., 1999; Zitko et al., 1998). This phenomenon was also reported by Bachour et al. (1998). They found that the concentration of highly chlorinated congeners was significantly higher on a lipid basis in muscle than in any of the other organs examined in humans. This
Table 3 Previously reported data for PCBs in human adipose tissue and liver (ng/g lipid) Country
Location
Year
Finland
Helsinki
1983 1982–1983
Finland Poland UK Italy Italy Sweden Belgium Belgium a
Turku Skierniewice Gdansk Wales Siena Siena
Antwerp Antwerp
On wet basis concentration.
1979 1990 1990–1991 1991–1992 1996–1997 1994 2000 2000–2001
Age (years) 15–91 19–95 27–91 35–68
21–83
19–77 5–76
Tissue
Range
Mean
Reference
Adipose Liver Adipose Adipose Adipose Adipose Adipose Adipose Adipose Liver Adipose Liver
80–1100 ND–9100 9.6–1996 750–1900 760–4700 257–2628
380 1100 504
Mussalo-Rauhamaa (1991)
931 1720 493a
Duarte-Davidson et al. (1994) Corsolini et al. (1995) Mariottini et al. (2000) Guvenius et al. (2001)
841 579
Covaci et al. (2002) Present study
364–686a 561–2343 475–2085 286–1802 70–1462
Smeds and Saukko (2001) Falandysz et al. (1994)
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situation can be explained as follows: for some predators, such as whales or seals, the daily intake of PCBs and DDTs is much higher than their elimination ability. In this case, liver, as a main organ for metabolism with a high blood flow rate, represents a depot organ for PCBs and DDTs. Bioaccumulation in terrestrial trophic chains was reported to be less accentuated than in the marine environment (Kumar et al., 2001). Being an omnivorous mammal, humans are exposed to PCBs primarily via low-level food contamination. Even in early investigations, market basket studies in the USA have shown a background PCB daily intake of 14 ng/kg/day (Gartrell et al., 1986). In the last decade, PCB contamination has decreased significantly, and data on PCB residues in human adipose tissues suggest that a steady state has been reached (Smeds and Saukko, 2001). Differing from liver, muscle tissue represents a comparatively inactive compartment with a low rate of metabolism. When the intake and elimination in humans has reached a steady state, the distribution of lipophilic compounds reaches an equilibrium and depends mostly on the transport mechanism. The concentrations of PCBs and DDTs in kidney and brain were significantly lower than in liver and muscle. Bachour et al. (1998) investigated 25 deceased males and females and found that the concentration of PCBs in brain was lower than in liver and muscle, especially for highly chlorinated congeners. Although a slightly lower concentration of SPCB was observed in the livers of children (samples 6L and 7L), the number of samples analyzed was too small to examine age- and genderspecific differences in PCB concentrations. While some data suggested an increase in concentrations with age (Robinson et al., 1990; Smeds and Saukko, 2001), other studies showed no such differences in PCB
serum
adipose
liver
concentrations with age (Kang et al., 1997). It seems that the relationship between the SPCB levels in human tissue is not as simple as we expected (Mariottini et al., 2000). In our study, subjects 4–6 were from the same family and therefore dietary differences could be neglected. While the ages of individual members were 5, 31, and 35 years, the total PCB concentrations in liver were similar for all three specimens (11.2, 14.4, and 12.7 ng/g wet wt, respectively). 3.2. Distribution of PCB congeners in human tissues The PCB congener patterns in different organs from individual subjects were similar (Table 2). The PCB pattern in our subjects and the previously reported distribution of PCB congeners in adipose tissue (Covaci et al., 2002) and serum (Covaci, 2002) are shown in Fig. 1. There were no significant differences between the distribution of congeners with different degrees of chlorination. The most abundant PCB congeners were PCB 153 and PCB 180, each of them accounting for approximately 30% and 25%, respectively, of the sum PCB in all samples. It is well know that the PCB pattern shifts from low-chlorinated congeners to higher chlorinated congeners when the organisms move from lower to higher trophic levels (Jones, 1988; Kumar et al., 2001). The higher concentrations of PCB 153 and PCB 180 (persistent congeners) and the absence of lowchlorinated congeners in individual subjects indicated that the principal source of contamination with PCBs was from diet and not from direct exposure. It is most likely that the contaminants passed several bioaccumulation steps before being ingested. Lanting et al. (1998) analyzed PCBs in adipose tissue, liver, and brain of nine fetuses that died in uteri. The
muscle
kidney
brain
40
% of sum 9 PCB congeners
35 30 25 20 15 10 5 0 PCB105
PCB118
PCB138
PCB153
PCB156
PCB170
PCB180
PCB187
PCB194
Fig. 1. Distribution of selected PCB congeners in human serum and different tissues (muscle, kidney, liver, and brain).
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medians (ranges) for the sum of PCBs 118, 138, 153, and 180 in the tissues were: adipose tissue 235 (97–768), liver 198 (67–362), and brain 50 (22–122) ng/g lipid. There were no differences in the distributions of congeners 118, 138, 153, and 180 between the three tissues and there was no statistically significant association between tissue PCB levels and gestational age (Lanting et al., 1998). We noticed that congeners with less than three chlorine atoms were not found in the samples. This phenomenon is different from those found in previous investigations. In 1991, the National Human Adipose Tissue Survey revealed that infants and children (age o14) had detectable levels of tri-CB but no nona-CBs. However, in the group aged 15–44 tri-CBs were not detectable but nona-CBs were (Phillips and Birchard, 1991). 3.3. Non-ortho PCB congeners Although non-ortho PCBs accounted for only 0.1% of SPCBs, these congeners are of great concern because of their dioxin-like toxicity. Similar to 2,3,7,8-TCDD, they bind to the cytosolic aryl hydrocarbon receptor to build a substrate–receptor complex that can enter the cell nucleus and interfere with gene expression. Thus, monitoring of toxic non-ortho PCBs is necessary for understanding their possible impact on human health (Corsolini et al., 1995). Data for non-ortho PCBs in studied human tissues are presented in Table 4. Normally, a decrease in the concentrations of the three non-ortho PCBs with an increasing degree of chlorination was found in water, soil, and sediment samples. PCB 77 was also abundant
173
in some technical mixtures (Erickson, 1997) and in animals which were situated at the bottom of the food chain. PCB 126 is abundant in some predators, such as birds, fox, and marine mammals, while PCB 169 is occasionally found in fish, birds, and some terrestrial animals, with a lower concentration relative to the other two congeners (Kumar et al., 2001). However, in our subjects this distribution was different from those found in other terrestrial mammals. In 10 of 18 tissue samples, the concentrations of PCB 169 were higher than those of PCB 126 (Table 4). This distribution of non-ortho PCBs in human tissue may have a time trend. Choi et al. (2002) investigated non-ortho PCB concentrations in Japanese human adipose tissue. The concentrations (mean7SD) of PCB 126 and PCB 169 in 1970–1971 were 2027146 and 50725 pg/g lipid wt, respectively, while in 2000 they were 72760 and 61725 pg/g lipid wt. Patterson et al. (1994) determined non-ortho PCBs in human adipose tissues from Atlanta. The mean concentrations of PCB 126 and PCB 169 were 102 and 67 pg/g lipid, respectively. This distribution might have been due to the fact that previously high PCB levels might have induced detoxifying hepatic enzymes (cytochrome P450) that lead to the degradation of less persistent congeners, such as PCB 77 and PCB 126 (Corsolini et al., 1995; Kannan et al., 1993; Patterson et al., 1994). The relative abundance of the three non-ortho PCB congeners in different tissues is similar in liver, muscle, and kidney. The mean concentration of PCB 77, PCB 126, and PCB 169 were 31.0, 56.6, and 59.1 pg/g lipid wt in liver samples. These values are lower than those
Table 4 Concentration of non-ortho PCBs in human tissues (pg/g wet weight) Subject no. 1
2
3
4 5 6 7 8 9 10 11 a b
Sample no.
Lipids (%)
PCB 77 (TEF:0.0001)
PCB 126 (TEF:0.1)
PCB 169 (TEF:0.01)
No-TEQa (pg/g wet wt)
Mo-TEQb (pg/g wet wt)
1B 1K 1L 1M 2K 2L 2M 3K 3L 3M 4L 5L 6L 7L 8L 9L 10L 11L
0.4 2.9 3.6 3.7 1.8 5.8 1.4 3.6 5.8 3.1 3.6 4.4 5.5 3.8 2.6 13.8 3.7 5.7
2.6 2.2 o2.0 2.2 o2.0 o2.0 o2.0 2.8 o2.0 o2.0 o2.0 o2.0 o2.0 o2.0 o2.0 2.4 o2.0 5.8
0.6 1.3 3.5 2.8 o0.4 1.1 o0.4 2.2 1.9 1.1 2.5 2.6 3.5 1.4 2.2 3.9 3.7 1.7
0.4 3.1 4.2 5.1 o0.4 2.1 0.5 3.2 6.3 3.3 1.3 0.8 0.8 o0.4 3.8 4.7 2.1 1.2
0.07 0.16 0.39 0.33 o0.04 0.13 o0.04 0.25 0.25 0.14 0.27 0.27 0.36 0.14 0.26 0.43 0.39 0.18
0.23 0.52 0.93 0.54 0.07 0.38 0.12 0.03 1.37 0.77 0.19 0.27 0.24 0.07 0.51 1.08 0.54 0.41
The No-TEQ values were calculated from the concentrations of non-ortho PCB congeners. The Mo-TEQ values were calculated from the concentrations of mono-ortho PCB congeners 105, 118, 156, and 189.
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previously reported in adipose tissue from Italy, Japan, Poland, and North America (Corsolini et al., 1995; Greve and Van Zoonen, 1990; Tanabe et al., 1989; Williams and Lebel, 1991). Toxic equivalent (TEQ) values based on the most recent toxic equivalency factors (TEFs) (van den Berg et al., 1998) were calculated to estimate the relative contributions of toxic potential from congeners know to cause induction in the detoxifying enzyme activities of the P450 family. They ranged from o0.04 to 0.43 and from 0.07 to 1.37 pg/g TEQ wet wt for non-ortho and mono-ortho PCBs, respectively (Table 4). The mean TEQ values were 0.28, 0.22, 0.21, and 0.07 pg/g wet wt in liver, muscle, kidney, and brain, respectively. The TEQ of the non-ortho PCBs was lower than that of mono-ortho congeners. TEQ values calculated for the three main mono-ortho PCBs (PCB 105, PCB 118, and PCB 156) were 0.51, 0.66, 0.31, and 0.23 pg/g wet wt in liver, muscle, kidney, and brain, respectively. The highest contribution was from PCB 156, which accounted for about 60% of the total PCB TEQ. 3.4. Enantiomeric composition of chiral PCBs and aHCH in human tissue The ERs for chiral PCBs and a-HCH investigated in this study are given in Table 5. The concentration of PCB 84 was very low in all samples. Due to the low signal-to-noise ratio and poor resolution, reliable quantification for PCB 84 atropisomer in the samples was impossible. In 3 of 18 samples (5L, 10L, and 11L), the enantiomeric ratio of a-HCH could be determined (ER=1.06, 1.09, and 1.02, respectively). Racemic or nearly racemic ratios for chiral PCBs were also found in
all muscle, kidney, and brain samples, ranging from 1.02 to 1.32 for PCB 95, 1.00 to 1.18 for PCB 149, and 0.88 to 1.09 for PCB 132. Although some papers suggested that enantioselective enrichment might occur in the brain (Vetter and Schurig, 1997), there was no enrichment of chiral PCBs in the brain sample. Most of the enantiomeric enrichment was found in liver samples. The mean (range) ERs in liver were 1.69 (1.04–2.97), 1.16 (0.99– 1.41), and 0.74 (0.48–0.97) for PCB 95, PCB 149, and PCB 132, respectively. The first attempt to determine the enantiomeric ratio of chiral PCB in human fluids was done by Glausch et al. (1994). Subsequent research has indicated that the second-eluted enantiomer of PCB 132 was enriched in human milk extracts (Glausch et al., 1995). Since the elution order of PCB 132 atropisomers on Chirasil-Dex was known (Haglund and Wiberg, 1996), (+)-PCB132 was the most abundant enantiomer in human milk. Because the same column (Chirasil-Dex) was used for chiral separation in our work, the results indicate that, similarly to milk, an enrichment of (+)-PCB 132 atropisomers was observed in human liver. All ERs of PCB 95 and PCB 149 were 41 or approximately equal to 1, while most of ERs for PCB 132 were o1 or close to 1. The elution order for PCB 95 and PCB 149 atropisomers on the Chirasil-Dex column might be reversed with that of PCB 132 atropisomers, but these results should be confirmed with pure PCB 95 and PCB 149 atropisomers. Relationships between the ERs of three chiral PCBs are shown in Figs. 2a and b and suggest that the bioselective metabolism of chiral PCBs has the same trend, although the ratio is different. The ERs for individual chiral PCBs in samples from 4L, 5L, 3.50
Subject no.
Sample no.
ER (a-HCH)
ER (PCB 95)
ER (PCB 149)
ER (PCB 132)
1.02 1.02 1.04 1.03 1.03 1.26 1.06 1.32 1.17 1.12 1.84 2.95 2.97 1.29 1.69 2.10 1.14 1.18
1.05 1.01 0.99 1.00 1.02 1.04 1.03 1.18 1.02 1.02 1.32 1.41 1.39 1.06 1.29 1.22 1.00 1.04
1.09 1.06 0.97 0.99 1.00 0.81 1.02 0.95 0.93 0.88 0.66 0.50 0.48 0.97 0.65 0.50 0.91 0.78
3.00
ER(PCB 95)
Table 5 Enantiomeric ratios of chiral PCBs and a-HCH in human tissues
2.50
r = 0.9433, p<0.001
2.00 1.50 1.00
2
3
4 5 6 7 8 9 10 11
1B 1K 1L 1M 2K 2L 2M 3K 3L 3M 4L 5L 6L 7L 8L 9L 10L 11L
1.06
1.09 1.02
0.50 0.50
1.00
1.50
2.00
2.50
1/ER(PCB 132)
(a) 1.50 1.40
ER(PCB 149)
1
1.30
r = 0.8674, p<0.001
1.20 1.10 1.00 0.90 0.80
(b)
1.00
1.20
1.40
1.60
1.80
2.00
2.20
1/ER(PCB 132)
Fig. 2. Relationships between ER of PCB 95 with 1/ER of PCB 132 (a) and ER of PCB 149 with of 1/ER 132 (b) in human tissues.
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and 6L (Table 1) were similar and showed that they were more depended on daily intake and not on age or sex.
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