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Biomonitoring Summary

Dioxin-Like Chemicals: Polychlorinated Dibenzo-p-dioxins, Polychlorinated Dibenzofurans, and Coplanar and Mono-ortho-substituted Polychlorinated Biphenyls

General Information

Polychlorinated dibenzo-p-dioxins and dibenzofurans are two similar classes of chlorinated aromatic chemicals that are produced as contaminants or by-products. They have no known commercial or natural use. Dioxins are produced primarily during the incineration or burning of waste; the bleaching processes used in pulp and paper mills; and the chemical syntheses of trichlorophenoxyacetic acid, hexachlorophene, vinyl chloride, trichlorophenol, and pentachlorophenol. Both the synthesis and heat-related degradation of polychlorinated biphenyls (PCBs) will produce dibenzofuran byproducts. Releases from industrial sources have decreased approximately 80% since the 1980s (U.S. EPA, 2004). Today, the largest release of these chemicals occurs as a result of the open burning of household and municipal trash, landfill fires, and agricultural and forest fires. When advanced analytical techniques are used, most soil and water samples will reveal trace amounts of polychlorinated dibenzo-p-dioxins and dibenzofurans.

The coplanar and mono-ortho-substituted PCBs are chlorinated aromatic hydrocarbon chemicals that belong to the general class PCBs which were once synthesized for use as heat-exchanger, transformer, and hydraulic fluids, and also used as additives to paints, oils, window caulking, and floor tiles. Production of PCBs peaked in the early 1970s and was banned in the United States after 1979. Together with the polychlorinated dioxins and furans, these two special classes of PCBs are often referred to as "dioxin-like" chemicals because they act in the body through a similar mechanism. Structural nomenclature is available at: https://www.epa.gov/oswer/riskassessment/pdf/1340-erasc-003.pdf. Commonly measured polychlorinated dibenzo-p-dioxins, polychlorinated dibenzofurans, and coplanar and mono-ortho-substituted PCBs are listed in the table.

Dioxin-like Chemicals Measured in the National Biomonitoring Program
Polychlorinated dibenzo-p-dioxins	CAS number
1,2,3,4,6,7,8-Heptachlorodibenzo-p-dioxin (HpCDD)	35822-46-9
1,2,3,4,7,8-Hexachlorodibenzo-p-dioxin (HxCDD)	39227-28-6
1,2,3,6,7,8-Hexachlorodibenzo-p-dioxin (HxCDD)	57653-85-7
1,2,3,7,8,9-Hexachlorodibenzo-p-dioxin (HxCDD)	19408-74-3
1,2,3,4,6,7,8,9-Octachlorodibenzo-p-dioxin (OCDD)	3268-87-9
1,2,3,7,8-Pentachlorodibenzo-p-dioxin (PeCDD)	40321-76-4
2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)	1746-01-6
Polychlorinated dibenzofurans	CAS number
1,2,3,4,6,7,8-Heptachlorodibenzofuran (HpCDF)	67562-39-4
1,2,3,4,7,8,9-Heptachlorodibenzofuran (HpCDF)	55673-89-7
1,2,3,4,7,8-Hexachlorodibenzofuran (HxCDF)	70648-26-9
1,2,3,6,7,8-Hexachlorodibenzofuran (HxCDF)	57117-44-9
1,2,3,7,8,9-Hexachlorodibenzofuran (HxCDF)	72918-21-9
2,3,4,6,7,8-Hexchlorodibenzofuran (HxCDF)	60851-34-5
1,2,3,4,6,7,8,9-Octachlorodibenzofuran (OCDF)	39001-02-0
1,2,3,7,8-Pentachlorodibenzofuran (PeCDF)	57117-41-6
2,3,4,7,8-Pentachlorodibenzofuran (PeCDF)	57117-31-4
2,3,7,8-Tetrachlorodibenzofuran (TCDF)	51207-31-9
Coplanar polychlorinated biphenyls (IUPAC number)	CAS number
3,3',4,4'-Tetrachlorobiphenyl (PCB 77)	32598-13-3
3,4,4',5-Tetrachlorobiphenyl (PCB 81)	70362-50-4
3,3',4,4',5-Pentachlorobiphenyl (PCB 126)	57465-28-8
3,3',4,4',5,5'-Hexachlorobiphenyl (PCB 169)	32774-16-6
Mono-ortho-substituted polychlorinated biphenyls (IUPAC number)	CAS number
2,3,3',4,4'-Pentachlorobiphenyl (PCB 105)	32598-14-4
2,3,4,4',5-Pentachlorobiphenyl (PCB 114)	74472-37-0
2,3',4,4',5-Pentachlorobiphenyl (PCB 118)	31508-00-6
2.3'4,4'5'-Pentachlorobiphenyl (PCB 123)	65510-44-3
2,3,3',4,4',5-Hexachlorobiphenyl (PCB 156)	38380-08-4
2,3,3',4,4',5'-Hexachlorobiphenyl (PCB 157)	69782-90-7
2,3',4,4',5,5'-Hexachlorobiphenyl (PCB 167)	52663-72-6
2,3,3',4,4',5,5'-Heptachlorobiphenyl (PCB 189)	39635-31-9

In the environment, these dioxin-like chemicals are persistent and usually occur as a mixture of congeners (i.e., compounds that differ by the numbers and positions of chlorine atoms attached to the dibenzo-p-dioxin, dibenzofuran, or biphenyl structures). The general population is exposed to low levels of polychlorinated dibenzo-p-dioxins and dibenzofurans primarily through ingestion of high-fat foods such as dairy products, eggs, and animal fats, and some fish and wildlife. Dioxin-like chemicals are measurable in U.S. meats and poultry (Hoffman et al., 2006) as a result of the accumulation of these substances in the food chain. Breast milk is a substantial source of exposure for infants (Beck et al., 1994; Lundqvist et al., 2006), though breast milk levels have been decreasing in recent years (Arisawa et al., 2005). The lesser chlorinated PCBs, including some dioxin-like PCBs, are more volatile. These PCBs can enter air of buildings containing joint sealants made with PCBs prior to 1980 and can increase background serum levels via inhalational exposure (Johansson et al., 2003; Kohler et al., 2005). Volatilization of PCBs from nearby hazardous waste sites may also contribute to human inhalational exposure. Exposure to high levels of these chemicals has occurred in the past as a result of industrial accidents (e.g., explosion of a factory in Seveso, Italy); the use of accidentally contaminated cooking oils (e.g., Yusho in Japan and Yucheng in Taiwan); the spraying of herbicides contaminated with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) (e.g., Agent Orange in Vietnam); and the burning of PCBs producing polychlorinated dibenzofurans (e.g., electrical transformer fires). Workplace exposures are infrequent today, but incineration plant workers and chemical synthesis workers can be exposed via inhalation and dust exposures. The dioxin-like chemicals are easily absorbed, tend to distribute into body fat, have limited metabolism, and are slowly eliminated from the body. Serum levels may be influenced by both past (stored in body fat) and recent exposures, though the current intakes for most people are now low. Half-lives of the dioxins and furans in the body vary from three to 19 years, with the half-life of TCDD estimated at around seven years (Geyer et al., 2002).

Because exposure to these chemicals includes a mixture of varying congeners, congener-specific effects are difficult to determine (Masuda, 2001; Masuda et al., 1998). However, these four groups of chemicals (polychlorinated dibenzo-p-dioxins, polychlorinated dibenzofurans, and the coplanar and mono-ortho-substituted PCBs) are considered to act through a similar mechanism to produce toxic effects. These dioxin-like effects are thought to result from interaction with the aryl hydrocarbon receptor (AhR), particularly in the induction of gene expression for cytochromes P450, CYP1A1 and CYP1A2. Dioxins and furans have a planar configuration and require four lateral chlorine atoms (2,3,7,8 positions) on the dibenzo-p-dioxin or dibenzofuran backbone to bind this receptor. The rank order of interaction with the AhR receptor by degree and position of chlorination is roughly similar for both the dioxin and furan series. The coplanar polychlorinated biphenyls (unsubstituted at ortho positions) and the mono-ortho-substituted polychlorinated biphenyls (which contain a chlorine atom at one of the ortho positions) can achieve a planar configuration and also interact with the AhR receptor. The variation in the effect on AhR among the dioxin-like chemicals is 10,000-fold, with TCDD and 1,2,3,7,8-pentachlorodibenzo-p-dioxin being the most potent. To compare potency, each of these congeners has been assigned a potency value relative to TCDD (toxic equivalency factor [TEF]). When each TEF is multiplied by the concentration of the congener, a toxic equivalency (TEQ) value is obtained. Thus, the dioxin-like toxicity contributed by each of the polychlorinated dibenzo-p-dioxins, dibenzofurans, and PCBs can then be compared. The sum of all congener TEQs in a specimen (total TEQ) can be used to compare dioxin-like activity among specimens. Many of the dioxin-like PCBs have lower potency but are found at higher concentrations than TCDD (Kang et al., 1997; Patterson et al., 1994, Van den Berg et al., 2006), so these less potent chemicals may still contribute substantially to the total TEQ.

In animal studies, TCDD and dioxin-like chemicals demonstrated many effects including: altered transcription of genes; induction of various enzymes; wasting syndrome; hepatotoxicity; altered immune function; testicular atrophy; altered thyroid function; chloracne; porphyria; neurotoxicity; teratogenicity; and carcinogenicity (EPA, 2004). Since animal species differ dramatically in sensitivity to these chemicals, it is difficult to predict human health effects although animal studies have provided support to observations of effects in human populations. Health effects of exposure to dioxin-like chemicals in people have been observed as a result of industrial or accidental exposures involving large doses of these chemicals. Chloracne, biochemical liver test abnormalities, elevated blood lipids, fetal injury, and porphyria cutanea tarda have been reported in episodes following high exposure. Developmental effects in humans are of concern, because congenital anomalies and intrauterine growth retardation were observed in children born to mothers who used rice-bran oil contaminated with very high levels of PCBs and polychlorinated dibenzofurans, an episode referred to as the Yucheng outbreak, which means oil disease in Chinese (Rogan et al, 1988). Even at so-called environmental levels of exposure, serum levels of primarily non-dioxin-like PCBs, and some dioxin-like chemicals, have been associated with altered psychomotor development in newborns and children (Arisawa et al., 2005; Jacobsen and Jacobsen, 1996; Koopman-Esseboom et al., 1996; Longnecker et al., 2003; Lundqvist et al., 2006; U.S. EPA, 2004).

Cross-sectional associations of type II diabetes or markers of insulin resistance with serum levels of TCDD, other dioxin-like chemicals, non-dioxin-like PCBs, and organochlorine pesticides have been reported in both highly exposed and environmentally exposed human populations, though some studies have not found an association (Calvert et al., 1999; Everett et al., 2007; Fierens et al., 2003; Fujiyoshi et al., 2006; Henriksen et al., 1997; Kang et al., 2006; Kern et al., 2004; Lee et al., 2006; Michalek et al., 1999, and 2003) and in vitro and in vivo animal studies have provided possible mechanistic plausibility. Immune effects of dioxin-like chemicals and non-dioxin-like PCBs have been reported in animal studies (Carpenter, 2006; U.S.EPA, 2004), but few or consistent effects in humans have been observed (Baccarelli et al., 2002; Halperin et al., 1998; Jung et al., 1998; IARC, 1997).

The dioxin-like chemicals appear to weakly mimic or interfere with the action of estrogen; for instance, dioxin-like chemicals may decrease the effect of estrogen through induction of its metabolism. This action contrasts with the non-dioxin-like PCBs and their metabolites, which may have direct estrogenic actions (Carpenter, 2006; Wang et al., 2006; Yoshida et al., 2005). Dioxin and other organochlorine chemicals have been shown to interfere with male and female reproductive development in experimental and wild animals, particularly during gestational exposure (Gao et al., 1999; Grey and Ostby, 1995; Roman et al., 1998; Sonne et al., 2006; Theobald et al., 1997). In studies of women with environmental or accidental exposures, associations between dioxin-like chemical exposures and various reproductive endpoints (Eskenazi et al., 2003; Lawson et al., 2004; Schnorr et al., 2001; Warner et al., 2004 and 2007) and endometriosis (Eskenazi et al., 2002; Fierens et al., 2003; Heilier et al., 2005; Hoffman et al., 2007) have been either absent or of unknown significance, although animal studies have demonstrated reproductive effects at high doses (Arisawa et al., 2005; U.S. EPA, 2004). In men, lowered levels of testosterone have been associated with environmental and occupational exposures to dioxin-like chemicals (Dhooge et al., 2006; Egeland et al., 1994; Gupta et al., 2006; Henriksen et al., 1996; Johnson et al., 2001; Sweeney et al., 1998), and gonadal atrophy and lowered testosterone levels have been observed in animal studies.

TCDD is classified separately by the IARC and NTP as a known human carcinogen. The U.S. EPA (2004) and IARC (1997) concluded that the aggregate evidence supports an association between high-dose TCDD exposure (e.g., encountered in contaminated occupational settings or massive unintentional releases) and increases in the all-cancer category (Steenland et al., 2004). The Institute of Medicine (2005) concluded that human epidemiologic evidence is sufficient for a positive association of herbicides contaminated with TCDD and an increased risk for non-Hodgkin's lymphoma, Hodgkin's lymphoma, chronic lymphocytic leukemia, and soft tissue sarcoma. Other individual polychlorinated dibenzo-p-dioxins and dibenzofurans have not been studied sufficiently for IARC to classify their human potential for carcinogenicity, although EPA considers these other chemicals as likely human carcinogens (U.S.EPA, 2004). Both IARC and NTP consider polychlorinated biphenyls as likely and probable human carcinogens. Additional information about external exposure (i.e., environmental levels) and health effects is available from ATSDR at https://www.atsdr.cdc.gov/toxprofiles/index.asp and from U.S. EPA at https://www.epa.gov/ncea/pdfs/dioxin/nas-review/.

Biomonitoring Information

Serum levels of the dioxin-like chemicals reflect accumulated exposure because these chemicals are stored in body fat and only slowly eliminated. Observed differences in the levels of these chemicals in people are due in part to differences in environmental exposure. For instance, people who ate dioxin-like chemical contaminated fish from the Great Lakes region had mean lipid adjusted serum concentrations of these chemicals that were several times higher than background values in the U.S. population (Anderson et al., 1998; Falk et al., 1999; Hanrahan et al., 1999; Turyk et al., 2006). Observed differences between people may also be due to longer periods of accumulation of these persistent chemicals. Several studies have shown that levels of the more highly chlorinated dioxins, furans, and PCBs in serum or fat will increase with the age of the population studied (Falk et al., 1999; Geyer et al., 2002; Kang et al., 1997; Luotamo et al., 1991; Patterson et al., 1986). Many of the dioxins, furans, and PCBs measured pooled serum samples obtained from a representative New Zealand population showed that levels trended upward with age (Bates et al., 2004). In a U.S. representative sample from NHANES 1999-2000, participants aged 20 years and older had higher levels than participants aged 12-19 years when levels at the higher percentiles of the more highly chlorinated congeners were compared (CDC, 2013). Similarly, the TEQ increased with age in an analysis of the NHANES 2003-2004 subsample (Patterson et al., 2009). Other factors also explain differences in levels observed between people. In a TEQ analysis of the NHANES 2001-2002 subsample, the total TEQ increased with age, was lower for Mexican Americans than non-Hispanic whites or blacks, and was higher in smokers than nonsmokers older than 60 years of age (Ferriby et al., 2007). Body mass is also a factor directly associated with increasing levels of some polychlorinated dibenzo-p-dioxins (Collins et al., 2007; Michalek et al., 1999). Gender is another predictor of levels of some dioxin-like chemicals. Compared with Japanese men, women had higher levels of octachlorodibenzo-p-dioxin, 1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxin, and 1,2,3,7,8,9-hexachlorodibenzo-p-dioxin, but men had higher levels of PCBs 169, 156, and 189 (Arisawa et al., 2003). In the NHANES 2001-2002 subsample, females had higher adjusted geometric mean levels than males for 1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxin, and 3,3',4,4',5-pentachlorobiphenyl (PCB 126). However, males had higher levels than females for 3,3',4,4',5,5'-hexachlorobiphenyl (PCB 169) (CDC, 2013).

The generally low lipid-adjusted levels observed in the U.S. representative NHANES subsamples of 1999-2000, 2001-2002, and 2003-2004 support the observation that human serum levels of polychlorinated dibenzo-p-dioxins, dibenzofurans, and PCBs have decreased by more than 80% since the 1980s (Aylward and Hays, 2002; Lorber, 2002; Patterson et al., 2009). Levels of some dioxin-like chemicals, such as the hexachlordibenzo-p-dioxins, were shown to decrease gradually from 1993 to 2003 in pooled samples from children in selected regions of Germany, whereas the hexachlordibenzofuran levels showed little change (Link et al., 2005). The levels of polychlorinated dibenzo-p-dioxins, dibenzofurans, and coplanar and mono-ortho-substituted biphenyls seen in the U.S. population were generally well below the levels associated with occupational or unintentional exposures that have produced health effects. There are no firmly established relationships between concentrations (mainly considering TCDD) and health effects in people. Observations following industrial and accidental exposures have suggested that acute exposures resulting in serum concentrations of about 800 pg/g of lipid might be necessary to induce clinical effects such as chloracne, although levels in the thousands of pg/g of lipid do not always produce this effect (Mocarelli et al., 1991). Such studies of clinical effects in people after large unintentional exposures have measured concentrations ranging from several hundred to the tens of thousands of pg/g of lipid of TCDD or equivalent (Eskenazi et al., 2004; Masuda, 2001; Masuda et al., 1998; Mocarelli et al., 1991). However, it has been suggested that background total TEQ for the general population are about 10-100 times the TEQ levels associated with a possible risk for adaptive or subclinical adverse effects (e.g., endocrine changes) (U.S.EPA, 2004).

In general, observations of the levels of dioxin-like chemicals across percentiles in the NHANES 2003-2004 subsample appeared roughly similar to previous NHANES surveys (CDC, 2013). For some chemicals, detection rates and percentile values will change over survey periods due to improvements in analytical methods and limitations of sample volume. In keeping with results from reports from Germany (Papke et al., 1998), New Zealand (Bates et al., 2004) and elsewhere, U.S. NHANES subsamples have shown that the highly chlorinated and laterally substituted congeners are detected most often (CDC, 2013, Patterson et al., 2009). The following listed dioxin-like chemicals were detectable in greater than 10% of the NHANES 2003-2004 subsample (those in bold letters were detectable in greater than 60% of the subsample). Many of these contribute to a significant portion of the total TEQ. The total TEQ at the 90th percentile of the U.S. population in NHANES 2003-2004 was 30.0 pg/g of lipid (Patterson et al. 2009).

1,2,3,4,6,7,8,9-octachlorodibenzo-p-dioxin
1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxin
1,2,3,6,7,8-hexachlorodibenzo-p-dioxin
1,2,3,7,8-pentachlorodibenzo-p-dioxin
2,3,7,8-tetachlordibenzo-p-dioxin
1,2,3,4,6,7,8-heptachlorodibenzofuran
1,2,3,4,7,8-hexachlorodibenzofuran
2,3,4,7,8-pentachlorodibenzofuran
coplanar PCBs 126 and 169
mono-ortho substituted PCBs 105, 118, 156, 157, 167, 189

Octachlorodibenzo-p-dioxin

Of the dioxins and furans measured in the U.S. representative subsamples of NHANES 1999-2000, 2001-2002, and 2003-2004, octachlorodibenzo-p-dioxin typically was present in the highest concentration, but contributed little to the TEQ, with the other commonly detected dioxin and furan congeners being more than eight-fold lower in concentration. Levels of octachlorodibenzo-p-dioxin that were similar to slightly higher than those in these NHANES subsamples were seen in a representative pooled sampling New Zealander residents aged 15 years and older obtained during 1997-1998 and also in a small convenience sample of German residents aged 18-71 years in 1996 (Bates et al., 2004; Papke et al., 1998; CDC, 2013). Similar levels were also found in 232 Belgian blood donors in 2000 (Debacker et al., 2007).

Hexachlorodibenzo-p-dioxins

The three major hexachlorodibenzo-p-dioxins are assigned equal TEF values, but the 1,2,3,6,7,8-hexachlorodibenzo-p-dioxin often demonstrated multifold higher concentrations than the other two hexachlorodibenzo-p-dioxins; about six times higher in the NHANES 2001-2002 subsample (CDC, 2013). The unadjusted geometric mean levels of 1,2,3,6,7,8-hexachlorodibenzo-p-dioxin in 2003-2004 and in 2001-2002 were 34.6 vs. 17.2 pg/g of lipid, respectively. The geometric mean levels of 1,2,3,6,7,8-hexachlorodibenzo-p-dioxin in the 2001-2002 subsample were slightly higher than levels in either the German or New Zealand study mentioned above (Bates et al., 2004; Papke et al., 1998). A convenience sample of Japanese men and women aged 20-76 years studied during 1996-1997 also showed lower median levels than levels in the NHANES 2001-2002 subsample (Arisawa et al., 2003; CDC, 2013).

1,2,3,7,8-Pentachlorodibenzo-p-dioxin

In prior NHANES surveys, 1,2,3,7,8-pentachlorodibenzo-p-dioxin concentrations were nearly 60-fold lower than octachlorodibenzo-p-dioxin levels (at the comparable percentiles) (CDC, 2013), but because of a 10,000-fold greater TEF (equal to that of TCDD), the contribution of 1,2,3,7,8-pentachlorodibenzo-p-dioxin to the total TEQ would be about 160 times greater than the octachlorodibenzo-p-dioxin. Levels of 1,2,3,7,8-pentachlorodibenzo-p-dioxin for the total population at the 95th percentile in the NHANES 2001-2002 and 2003-2004 subsamples were 15.8 pg/g and 11.0 pg/g lipid, respectively. In 1996, a convenience sample of German residents aged 18-71 years showed that levels of 1,2,3,7,8-pentachlorodibenzo-p-dioxin at the 95th percentile were 9.9 pg/g lipid (Papke et al., 1998). The 95th percentile of a group of workers with distant past trichlorophenol exposure was about twice as high as the 95th percentile for adults in NHANES 2001-2002 (CDC, 2013; Collins et al., 2006).

2,3,7,8-Tetrachlorodibenzo-p-dioxin

TCDD is considered the most potent of the dioxin-like chemicals and environmental exposure usually results in very low serum concentrations. In the NHANES 2003-2004 subsample, the 95th percentile for the total population (12 years and older) was 5.2 picograms/gram (pg/g) of lipid. In 1996, the 95th percentile for lipid-adjusted serum TCDD levels in 139 Germans aged 18-71 years was 4.3 pg/g of lipid, with that percentile comprising mainly older individuals (Papke, 1998). In contrast, the most highly exposed females following the Seveso, Italy, factory explosion had median lipid adjusted levels of 272 pg/g lipid in 1976 (Eskenazi et al., 2004). TCDD levels in chemical plant workers with higher exposures have ranged as high as 2,000 pg/g lipid (IARC, 1997). Median serum TCDD levels measured in chemical production workers 15 years after workplace exposure ended were 68 pg/g of lipid (Calvert et al., 1996; Calvert et al., 1999). TCDD levels in the U.S. general population were also lower than workers with past trichlorophenol exposure (Collins et al., 2006) and lower than Vietnam veterans 20 years after duty-related exposure to Agent Orange (median serum TCDD concentration was 12.2 pg/g of lipid) (Henriksen et al., 1997).

Polychlorinated dibenzofurans

Of the polychlorinated dibenzofurans, the following could be characterized at the 95th percentiles (or lower) in the NHANES 1999-2000, 2001-2002 and 2003-2004 subsamples: 1,2,3,4,6,7,8-heptachlorodibenzofuran, 1,2,3,4,7,8-hexachlorodibenzofuran, 1,2,3,6,7,8-hexachlorodibenzofuran, and 2,3,4,7,8-pentachlorodibenzofuran. Generally, these levels are similar to other large population studies. In 237 workers with past exposure to trichlorophenol, where little polychlorinated dibenzofuran exposure would be expected, higher percentiles values were similar to a referent population and to the NHANES 1999-2000 and 2001-2002 subsamples (Collins et al., 2007; CDC, 2013). In 232 Belgian blood donors from the year 2000, the geometric mean level of 1,2,3,4,6,7,8-heptachlorodibenzofuran was several times lower than the geometric mean value in the NHANES 2001-2002 subsample of adults and the other dibenzofurans examined in the Belgian donors were lower than the limits of detection in NHANES 2000-2001 (CDC, 2013; Debacker et al., 2007). In Yucheng rice oil contamination victims when examined 15 years after their exposure, levels of the polychlorinated dibenzofurans were still hundreds of times higher than in levels for the U.S. population observed in the NHANES subsamples (Hsu et al., 2005).

Coplanar PCBs

The coplanar PCBs typically contribute less than about 15% to the total TEQ in the U.S. population (Ferriby et al., 2007). In the NHANES 2001-2002 subsample, the geometric mean levels of PCBs 126 and 169 for adults aged 20 years and older were similar or slightly lower than those reported from a representative pooled sample of New Zealanders in 1996-1997 (Bates et al., 2004; CDC, 2013) and from a smaller sample of non-occupationally exposed men and women aged 20-76 years in Japan in 1999 (Arisawa et al., 2003). Higher levels of these PCBs have been reported for persons consuming sport fish caught in the Great Lakes region (Turyk et al., 2006). In 311 residents of northern Italy, serum PCB 126 and 169 were not detectable, though other PCBs tended to be higher than in the recent NHANES subsamples (Apostoli et al., 2005; CDC, 2013).

Mono-ortho-substituted PCBs

Of the mono-ortho-substituted PCB congeners, the most frequently detected in general population studies are PCBs 118 and 156. Of these, PCB 118 levels were higher than levels of PCB 156 in the NHANES 1999-2000, 2001-2002, and 2003-2004 subsamples, although PCB 156 contributes more to the TEQ because its TEF is five-fold greater than the TEF of PCB 118. Although these PCBs are relatively less potent (i.e., lower TEFs), their contribution to the total TEQ in the U.S. population is about 25% (Ferriby et al., 2007) since they are present in much higher concentrations than are the coplanar PCBs, dioxins, and furans. In a convenience sample of the U.S. population in 1988 (Patterson et al., 1994), levels of PCB 118 were five-fold higher than in the NHANES 1999-2002 subsamples (CDC, 2013). Comparable levels of PCB 156 levels in NHANES 1999-2000 were slightly lower than those reported for a Canadian population study in 1994 (Longnecker et al., 2000). In a referent population of 311 residents in northern Italy during 2001-2003, the 95th percentile levels of PCB 156 and PCB 118 were two to three times higher than for the NHANES 1999-2002 subsamples (Apostoli et al., 2005; CDC, 2013). Levels of PCB 156 and PCB 118 were slightly higher in a Swedish study of 150 men than in the NHANES 1999-2000 subsample, possibly due to higher fish intake in the Swedish population (Glynn et al., 2000; CDC, 2013). However, in fish-consuming Japanese men and women studied during 1996-1997, PCB 118 levels at the 75th percentile were similar to levels in the NHANES 2001-2002 subsample (Arisawa et al., 2003).

Finding a measurable amount of one or more of the polychlorinated dibenzo-p-dioxins, dibenzofurans, coplanar or mono-ortho-substituted biphenyls in serum does not mean that the level of one or more of these chemicals causes an adverse health effect. Biomonitoring studies of serum polychlorinated dibenzo-p-dioxins, dibenzofurans, coplanar or mono-ortho-substituted biphenyls provide physicians and public health officials with reference values so that they can determine whether or not people have been exposed to higher levels of polychlorinated dibenzo-p-dioxins, dibenzofurans, coplanar or mono-ortho-substituted biphenyls than levels found in the general population. Biomonitoring data can also help scientists plan and conduct research on exposure and health effects.

References

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