|Moderator:Jouni (see all)|
9. TOXIC EQUIVALENCY FACTORS (TEF) FOR DIOXIN AND RELATED COMPOUNDS
Previous risk assessments of dioxin and dioxin-like compounds from around the world have employed the Toxic Equivalency Factor (TEF) methodology. This method is also used throughout EPA’s dioxin reassessment. This chapter has been added to the EPA’s dioxin reassessment effort to address questions raised by the Agency’s Science Advisory Board (SAB) in 1995. In its Report to the Administrator (U.S. EPA, 1995), the Committee said it "supports EPA’s use of Toxic Equivalencies for exposure analysis..." However, the SAB suggested that, as the toxic equivalent (TEQ) approach was a critical component of risk assessment for dioxin and related compounds, the Agency should be explicit in its description of the history and application of the process and go beyond reliance on the Agency’s published reference documents on the subject (U.S. EPA, 1987; 1989; 1991) to discuss issues raised in review and comment on this approach. Significant additional literature is now available on the subject, and this chapter provides the reader with a summary which is up-to-date through 2000. Future research will be needed to address uncertainties inherent in the current approach. The World Health Organization (WHO) has suggested that the TEQ scheme be reevaluated every 5 years and that TEFs and their application to risk assessment be re-analyzed to account for emerging scientific information (van den Berg et al., 1998).
WHO TEF workgroup 2005
The WHO 1998 TEF was set at 1.0, which is above the 90th percentile of the REP distribution of 12 in vivo studies. New studies indicate an REP between 0.1 and 1.0 for this compound (Fattore et al., 2000; Johnson et al., 2000; Simanainen et al., 2002). The vitamin A and tumor promotion studies provide REPs of 0.7 and 1 (Fattore et al., 2000; Waern et al., 1991). Results from acute toxicity studies result in REPs closer to 0.5, but, in general, REPs increase in subchronic toxicity studies (Haws et al., 2006). Therefore, the consensus WHO 2005 TEF value remained at 1.
The WHO 1998 TEF was set at 0.1, which is around the 80th percentile of the REP distribution of five in vivo studies. One new study determined REPs ranging from 0.04 to 0.12 (Gao et al., 1999), while two other recent studies observed REPs between 0.06 and 0.4 (Simanainen et al., 2002; Takagi et al., 2003). Very little data indicate that the TEF value should be changed to either 0.3 or 0.03. Therefore, it was decided to keep the WHO 2005 TEF value at 0.1.
The WHO 1998 TEF was set at 0.1. No in vivo studies are available for this HxCDD isomer. This TEF value is above the 75th percentile of the REP distribution of four in vitro studies (Haws et al., 2006). A more recent in vitro study (Bols, 1997) supports this TEF value, and therefore, no change for the WHO 2005 TEF was decided.
Similar to the above previous hexa-isomers, the WHO 1998 TEF was set at 0.1. It was noted that very little in vivo data are available with a recent study giving an REP of 0.029 (Takagi et al., 2003). In addition, four in vitro studies produced REPs up to 0.07 (Lipp et al., 1992; Schrenk et al., 1991), which is above the 75th percentile of the distribution. The expert panel considered decreasing the TEF value to 0.03 but decided that there was not enough data to support such a change. In vitro data were observed to be consistent between HxCDD isomers. In view of the homology between the HxCDD isomers, it was therefore decided to retain the old value of 0.1 as the WHO 2005 TEF value.
The WHO 1998 TEF was set at 0.01, which is at the 50th percentile of the REP distribution range of four in vivo studies. New studies again point toward an REP of 0.01 for this congener (Simanainen et al., 2002; Viluksela et al., 1994, 1997a,b). An earlier tumor promotion study also indicated a similar REP (Schrenk et al., 1994). It was also discussed whether or not the available information from the important studies mentioned above would be sufficient to increase the TEF to 0.03, which is well above the 90th percentile of the REP distribution. This suggestion was rejected by the expert panel. It was decided to retain the WHO 2005 TEF value as 0.01.
The WHO 1998 TEF was set at 0.0001, which is well outside the 10th percentile of the range of in vivo and in vitro REP values (Haws et al., 2006). At present, the only in vivo REPs meeting the stringent conditions of the database (Haws et al., 2006) are based on one study that was reported in two different papers using different endpoints (Fattore et al., 2000; Wermelinger et al., 1990). It was discussed whether or not this TEF should be increased to bring it in line with the results of the subchronic toxicity study (Fattore et al., 2000; Wermelinger et al., 1990). The new in vivo REP data from Fattore et al. (2000) were evaluated, and these would support a TEF greater than 0.0001. One concern that was expressed within the expert panel was that the animals used in a more recent publication (Fattore et al., 2000) were the same animals used in an earlier study (Wermelinger et al., 1990), and this octachlorodibenzo-p-dioxin (OCDD) was reported to be contaminated with other more potent 2,3,7,8–substituted congeners such as 2,3,4,7,8-PeCDF. Using the NTP data now available for 2,3,4,7,8-PeCDF (Walker et al., 2005), it was calculated that the reported contamination of 2.5 ppm (pg/lg) 2,3,4,7,8-PeCDF was not of toxicological relevance for the results (Calculation of 2,3,4,7,8-PeCDF contamination in OCDD Fattore et al. study . 2.5 ppm = 2.5 pg/lg OCDD. Highest OCDD dose 800 ppb = 800 ng/g = 0.8 lg/g feed. At this dose level, the PeCDF dose must have been 2.5 pg PeCDF/0.8 lg OCDD/g feed, which is equivalent with 2 pg PeCDF/g feed. Assuming a rat of 200 g with 20 g feed per day, the PeCDF dose must have been 40 pg PeCDF/200 g rat, which is equivalent with 200 pg PeCDF/kg/day or 0.2 ng PeCDF/kg/ day. This dose is two orders of magnitude lower than the lowest dose (20 ng PeCDF/kg/day) used in the National Toxicology Program and well below the no observed effect level (NOEL) of all endpoints that were looked at.). Overall, it was concluded that there is very limited in vivo information available with only one subchronic toxicity study (Fattore et al., 2000; Wermelinger et al., 1990). The expert panel decided that the information provided by both in vivo studies derived from only one experiment did not provide a solid basis to increase the TEF value for this compound to 0.001, but the combined information from in vivo and in vitro data (Haws et al., 2006) did justify a raise in TEF value. Therefore, it was decided to increase the WHO 2005 TEF value to 0.0003. The expert panel is aware of the implications that the increase in this WHO TEF value for OCDD might have from a regulatory and risk management point of view. However, with respect to the high concentrations of OCDD in some environmental matrices, a number of critical remarks regarding the inappropriate use of the present WHO TEFs are made in the section on the use of TEQ for abiotic environmental matrices.
The WHO 1998 TEF was set at 0.1. This value is at the 75th percentile of the REP distribution of nine in vivo studies for this compound (Haws et al., 2006). Only one new study has been reported (Takagi et al., 2003), and an REP of 0.07 was found for increased cleft palate formation, which is close to the TEF of 0.1. Consequently, it was decided that the WHO 2005 TEF should remain at 0.1.
The WHO 1998 TEF was set at 0.05, which is within the 50th and 75th percentile of the REP distribution of eight in vivo studies. A new study by Fattore et al. (2000) found an REP of 0.01 for effects on hepatic vitamin A reduction, but a study by Takagi et al. (2003) reported an REP of 0.045 for cleft palate. The majority of the in vivo studies report an REP value below 0.1, but many relevant studies have REPs above 0.01. Therefore, it was decided that the 2005 WHO TEF should become 0.03.
The WHO 1998 TEF was set at 0.5, which is well above the 75th percentile of the REP distribution of eight in vivo studies. Results from the long-term NTP study in female Sprague- Dawley rats using many different endpoints are now available to evaluate this earlier TEF value more closely. The REPs for neoplastic endpoints from the NTP study (Walker et al., 2005) are around 0.2–0.3, while nonneoplastic endpoints have REPs that range from 0.7 to 1.1. An earlier subchronic study by Pluess et al. (1998) pointed toward an REP of 0.4. More recent studies using hepatic vitamin A reduction and immunological effects as endpoints also point toward a TEF below 0.5 (Fattore et al., 2000; Johnson et al., 2000). In view of this new information, it was decided by consensus of the expert panel to change the WHO 2005 TEF to 0.3.
The WHO 1998 TEF was set at 0.1, which is above the 75th percentile of the REP distribution of six in vivo studies. No new in vivo studies have been published since 1997, and in view of the limited data, there was no reason to change this value. Thus, the WHO 2005 TEF value remains 0.1. 1,2,3,6,7,8-HxCDF The WHO 1998 TEF was also set at 0.1, and this value is above the 75th percentile of the distribution of three in vivo REPs, and when the results of in vitro and in vivo studies with the PCDF are combined, REP values lie within the 50th and 75th percentile. A new study reported an REP of 0.03 for hepatic vitamin A reduction (Fattore et al., 2000). However, the animals analyzed were from an earlier study from which an REP of 0.1 for subchronic toxicity was reported (Pluess et al., 1998). In view of the limited number of studies available and the fact that WHO 1998 TEFs of 0.1 for most HxCDFs were all around the 50th to 75th percentile (Haws et al., 2006), the expert panel decided not to discriminate between TEF values for these congeners. As a result, the 2005 WHO TEF remains at 0.1.
The WHO 1998 TEF for this HxCDF was set at 0.1. There are no in vivo results, and only two earlier in vitro studies for this congener with REPs of 0.2 and 0.1 (Brown, 2001; Tysklind et al., 1994) supporting the 0.1 TEF value similar to the other HxCDFs. Consequently, the 2005 WHO TEF remains as 0.1.
The WHO 1998 TEF value is 0.1, and it is around the 50th percentile of the REP distribution range of the combined in vivo and five in vitro studies (Haws et al., 2006). Most in vitro studies suggest a TEF value slightly above 0.1 (Bandiera et al., 1984; Brown, 2001; Mason et al., 1987; Tysklind et al., 1994). There is only one in vivo study for this hexa-isomer indicating REPs for different endpoints ranging from 0.02 to 0.1. Given this weak and limited REP database and approximate similarities in responses for this and certain other HxCDFs, there was consensus in the expert panel to retain the 2005 WHO TEF at 0.1.
1,2,3,4,6,7,8- and 1,2,3,4,7,8,9-HpCDFs
The WHO 1998 TEFs for both HpCDFs were set at 0.01. Since 1997, there are no new in vivo studies published. Only two in vitro studies have been published (Brown, 2001; Tysklind et al., 1994) reporting REPs, respectively, of 0.02 and 0.3 for 1,2,3,4,6,7,8-HpCDF and 0.04 and 0.02 for 1,2,3,4,7,8,9-HpCDF. Although these in vitro results do suggest a slightly higher TEF than 0.01, the expert panel thought that there was too much uncertainty in this limited database to raise the TEF. In addition, it was expected that in vivo, there would be low absorption of these HpCDFs from the gastrointestinal tract, thereby reducing their relative potency below that of the in vitro REPs. Based on these arguments, it was decided that the WHO 2005 TEFs would remain the same for both isomers, 0.01.
The WHO 1998 TEF value of 0.0001 is within the 50th and 75th percentile of the REP distribution range of three in vivo studies, but when these data are combined with in vitro results, it falls below the 50th percentile (Haws et al., 2006). The recent study by Fattore et al. (2000) using the same animals as Wermelinger et al. (1990) indicate an REP for octachlorodibenzofuran (OCDF) greater than 0.0001 based on hepatic vitamin A reductions. Some earlier in vivo studies also indicated an REP higher than the WHO 1998 TEF (DeVito et al., 1998; van Birgelen et al., 1996a; Waern, 1995). As with OCDD, there was originally concern among the expert panel about impurities with 2,3,7,8-chlorine–substituted congeners (Fattore et al., 2000;Wermelinger et al., 1990), but calculations indicated that the reported contamination with 1,2,3,4,6,7,8- HpCDF was of no toxicological concern. When the limited number of in vivo and in vitro REPs (< 10) are reviewed, REPs range from 4 3 10E-6 to 0.0028 with a 50th percentile of 0.0007 (Haws et al., 2006). As with OCDD, the expert panel decided that the limited in vivo information available would not warrant a factor of 10 increase of the WHO 1998 TEF value, but increasing the WHO 2005 TEF value to 0.0003 is appropriate in view of some of the higher in vivo REPs reported. This would also be in line with comparable REP values obtained in a recent study including both OCDD and OCDF (Fattore et al., 2000).
The WHO 1998 TEF value of 0.0001 is just below the 75th percentile in a very nonhomogenous distribution of six in vivo REPs. The available subchronic toxicity studies are all around the 75th percentile (Chu et al., 1995; Hakansson et al., 1994). Immunotoxicological studies with mice were given less weight (Harper et al., 1995; Mayura et al., 1993) because these were acute studies involving the ip route of exposure, and no information on purity was provided. It was decided by the expert panel that the subchronic study was still the most representative (Chu et al., 1995). As a consequence, the WHO 2005 TEF value remained at 0.0001.
The WHO 1998 TEF value was 0.0001. PCB 81 has been observed in wildlife and human milk (Kumar et al., 2001), confirming the validity of inclusion of this PCB in the TEF scheme. There are no new in vivo data for this PCB congener. Older in vivo data were excluded because these involved single-dose studies from which the expert panel believed that no reliable REP value could be determined. Various in vitro studies with human hepatoma HepG2 cells and monkey hepatocytes indicate that PCB 81 is more potent than PCB 77 (Brown, 2001; Pang et al., 1999; van der Burght et al., 1999; Zeiger et al., 2001). Based on the in vitro REP distribution, it is noticeable that the WHO 1998 TEF is located at the very low end of the REP distribution range (Haws et al., 2006). Thus, based on the information that PCB 81 is more potent in vitro and more persistent than PCB 77, the expert panel decided to raise the WHO 2005 TEF value to 0.0003. However, the expert panel expressed its low confidence in the PCB 81 REP database because it lacks in vivo REP data.
The WHO 1998 TEF was set at 0.1, which is at the median of the REP distribution range of 20 in vivo studies. This 1998 TEF value was mainly driven by the tumor promotion study with this compound (Hemming et al., 1995). New in vivo studies from the NTP covering many endpoints (Johnson et al., 2000; Walker et al., 2005) support the TEF of 0.1. In rat studies, the expert panel recognized the tight range of REPs for this congener (Haws et al., 2006), which supports the use of PCB 126 as reference compound with a TEF of 0.1 when comparing rat studies. Information from mice studies and some human in vitro systems (especially for enzyme induction) suggest that the REP for PCB 126 might be lower than 0.1 (Birnbaum and DeVito, 1995; DeVito et al., 2000; Harper et al., 1993; van Birgelen et al., 1996a; van Duursen et al., 2003; Zeiger et al., 2001). Clearly, more information is necessary regarding this issue. Although concern was expressed about interspecies variability in REPs, the expert panel considered the present information too limited to make a decision other than to retain 0.1 as the WHO 2005 TEF.
The WHO 1998 TEF was set at 0.01, which is below the median in the REP distribution range of seven in vivo studies. The 1998 TEF was mainly driven by a 4-week repeated dose mouse study measuring enzyme induction and generating an REP of less than 0.001 (DeVito et al., 1998). On the other hand, REPs from several other in vivo studies ranged from less than 0.01 to 0.7 (Harper et al., 1993; Parkinson et al., 1981; Yoshimura et al., 1979). Thus, large differences in REPs have been observed for PCB 169 between both species and endpoints. In view of the fact that the WHO 1998 TEF was also below the median of the in vivo REP distribution (Haws et al., 2006), the expert panel decided that it was appropriate to raise the TEF between the 50th and 75th percentile (see Figure 1). Nevertheless, many single-dose studies were observed to have significantly higher REPs (around 0.1) than those observed in a 13-week study. In view of these significant differences between single- and multiple-dose studies, the expert panel judged that the WHO 2005 TEF for PCB 169 of 0.03 would be more appropriate than a potentially overly conservative REP of 0.1.
Mono-Ortho–Substituted PCBs 105, 114, 118, 123, 156, 157, 167, and 189
The WHO 1998 TEF values for the mono-ortho PCBs ranged from 0.00001 to 0.0005. A major issue with the REP values for the different mono-ortho PCBs is that they span four to five orders of magnitude, depending on the congener. In Figure 3, this wide variation in REPs is illustrated. Even if only in vivo studies are considered, the 90% distribution range is extremely large (see Figure 1). This great variation in REP values was of serious concern to the expert panel. The panel considers possible, inconsistent, and low level contamination of the mono-ortho PCBs with more potent dioxin-like compounds to play, at least in part, a role in causing this large variation. De Vito (2003) found that less than 1% contamination of PCB 77 by PCB 126 significantly impacted the apparent REP of PCB 77. Shortly before the WHO 2005 TEF reevaluation meeting, two laboratories of panel members performed a number of in vitro experiments in an attempt to elucidate the possible impact of impurities on REPs for the mono-ortho PCBs (Peters et al., 2006). This study showed that after being purified on charcoal, the mono-ortho PCBs 105, 118, 156, and 167 did not cause AhR-mediated activation and CYP1A1 induction in two genetically modified rodent hepatoma CAFLUX cell lines at concentrations that would generally justify an REP larger than 0.0001. Based on the combined information, the expert panel expressed low confidence in the higher REP values for certain mono-ortho PCBs. It was concluded that the unusually wide variability of REP values for monoortho PCBs can probably be explained by the occurrence of impurities with 2,3,7,8-chlorine–substituted PCDDs and PCDFs or PCB 126. As the occurrence of these impurities clearly depends on the route of synthesis and the degree of cleanup, it was not possible to make a general statement about how it occurs in all cases. It was concluded that for future studies with mono-ortho PCBs or any other weak AhR agonists, a purity of > 99% is clearly not sufficient to establish a reliable REP. The expert panel compiled Figure 3 to make a decision on the TEF values of the mono-ortho PCBs, acknowledging the impurity issue and that the earlier decision scheme with ≥ 75th percentile (Fig. 2) was not appropriate. In this case, the most environmentally relevant mono-ortho PCBs are 105, 118, and 156, and it was decided to use the medians of the REP distribution range of these PCB congeners as a guide. This resulted in a recommended TEF of 0.00003 for these three mono-ortho PCBs. A differentiation for all other remaining mono-ortho PCBs was considered not feasible by the expert panel due to the lack of sufficient experimental data. Consequently, the recommended WHO 2005 TEF for all mono-ortho PCBs is 0.00003.
Other compounds discussed for possible inclusion in the TEF scheme
PCB 37 is commonly found in the environment (Hansen, 1998). It has also been detected in edible fish species at levels comparable with PCB 77 and 81 (Sapozhnikova et al., 2004). In seals, it has been measured in relatively high levels, indicating possible bioaccumulative properties in the food chain (Addison et al., 1999). It has also been found in human milk (Hansen, 1998). In an in vitro study with the human MCF-7 breast carcinoma and HepG2 hepatoma cell lines, no induction of CYP1A1 or 1B1 could be found. However, PCB 37 was found to be a significant catalytic inhibitor of both CYP activities (Pang et al., 1999). In view of the above information, there is a clear need for more in vivo and in vitro information to decide if this PCB needs to be included in the TEF scheme.
Polybrominated Dibenzo-p-Dioxins and Polybrominated Dibenzofurans
Both in vitro and in vivo studies have shown that polybrominated dibenzo-p-dioxins (PBDDs) and polybrominated dibenzofurans (PBDFs) have AhR agonist properties and cause dioxin-like effects (Birnbaum et al., 2003; Mason et al., 1987). Emerging data from Japan and the Baltic Sea indicate that PBDDs and PBDFs can be found in sediment, mussels, and higher trophic species like the cormorant (Choi et al., 2003; Malmvarn et al., 2005; Takigami et al., 2005; Watanabe et al., 2004). In addition, there is limited recent information showing that these compounds are found in human milk and adipose tissue at levels that can contribute significantly to the total amount of TEQ (Choi et al., 2003; Kotz et al., 2005; Ohta et al., 2005). It appears that environmental levels might be significantly lower than those of the PCDDs, PCDFs, and PCBs already in the TEF scheme. However, a better exposure assessment especially with regard to humans is needed. If the presence of PBDDs and PBDFs in human food as well as in people is more extensively demonstrated, there would be a clear FIG. 3. Distribution of REP values for the different mono-ortho PCBs based on AhR-mediated effects. need for assigning TEFs to these compounds. At present, it is unclear to what extent the ongoing use of brominated flame retardants, especially polybrominated diphenylethers (PBDEs), could lead to an increase in human and environmental exposure to PBDDs and PBDFs. Therefore, it is recommended by the expert panel to perform a more thorough exposure analysis for humans. In addition, it was concluded that among all compounds proposed in this paragraph for development of WHO TEFs, the PBDDs and PBDFs should be given high priority. More REP studies on PBDDs and PBDFs are urgently needed.
Mixed Halogenated Dibenzo-p-Dioxins and Mixed Halogenated Dibenzofurans
Due to the extremely high number of congeners, analysis of mixed halogenated dibenzo-p-dioxins (PXCDDs) and mixed halogenated dibenzofurans (PXCDFs) is still a major problem. Very little is known about the possible relevance of these compounds for human exposure (Birnbaum et al., 2003). If the mixed halogenated (bromine- and chlorine-substituted) dioxins and dibenzofurans are indeed detected in humans and their food, these should definitely be considered for inclusion in the TEF scheme. Early in vitro studies suggest that these compounds follow the same structure-activity rules as the PCDDs and PCDFs (Behnisch et al., 2001; Mason et al., 1987; Weber and Greim, 1997).
It has been suggested that hexachlorobenzene (HCB) fulfills the criteria for inclusion in the TEF concept (van Birgelen, 1998), although arguments for doing so have been criticized (Pohl et al., 2001; Schwab, 1999; Vos, 2000). HCB has mixed inducer properties in analogy with the mono-ortho PCBs. Before inclusion in the TEF concept is considered, it should be confirmed that highly purified HCB has indeed AhR agonistic properties, as contamination of HCB with PCDDs and PCDFs has been reported (Goldstein, 1979) (Analysis of HCB done for the U.K. Medical Research Council indicated levels of 16,000 ng OCDD/g, 6000 ng OCDF/g, 1000 ng HpCDF/g, and 88 ng TCDD/g in HCB of high chemical quality [M. Rose, personal communication].). Thus, results from earlier HCB studies could have an impurity problem similar to that observed for the mono-ortho PCBs. Priority should thus be given to confirm the compound’s dioxin-like properties using highly purified HCB with measured absence of 2,3,7,8-chlorine–substituted dioxins and dibenzofurans or dioxin-like PCBs.
Polychlorinated Naphthalenes and Polybrominated Naphthalenes
Based on recent published data, there was agreement by the expert panel that these compounds definitely should be considered for inclusion in the TEF concept as polychlorinated naphthalenes (PCNs) are actually reported in food and humans (Domingo et al., 2003; Falandysz, 2003; Hayward, 1998; Lunden and Noren, 1998; Weistrand and Noren, 1998; Williams et al., 1993). Earlier in vivo studies demonstrated that PCNs and polybrominated naphthalenes (PBNs) were able to induce dioxin-like effects, such as cleft palate and hydronephrosis (Birnbaum et al., 1983; McKinney and McConnell, 1982; Miller and Birnbaum, 1986). Further arguments for inclusion would be that multiple PCN and PBN congeners have distinct in vitro AhR activities that show analogy with PCDDs and PCDFs but are less potent (Behnisch et al., 2003; Blankenship et al., 2000; Darnerud, 2003; Robertson et al., 1982, 1984; Villeneuve et al., 2000). However, as with mono-ortho PCBs and HCBs, the possible impurity issue should be addressed thoroughly before inclusion in the TEF concept is decided.
Certain polybrominated biphenyls (PBBs) have been reported to be AhR active in both in vitro and in vivo experiments (Darnerud, 2003; Robertson et al., 1982, 1984). It was noted by the expert panel that some human background exposure to PBBs is still occurring. However, human exposure data outside the ‘‘Michigan episode’’ are surprisingly scarce. Recently, PBB exposure has been reported in bird species at the top of the food chain from Japan and Norway (Herzke et al., 2005; Lindberg et al., 2004; Watanabe et al., 2004). This occurrence in top predator wildlife species also stresses the need to further identify present human background exposure to PBBs. Thus, based on the AhR mechanism of action, inclusion of certain PBB congeners in the TEF scheme is appropriate, but further human exposure analysis should identify the possible relevance of PBBs to the total TEQ.
The expert panel accepted that PBDEs by themselves do not have AhR agonist properties and should not be included in the TEF concept (Chen and Bunce, 2003; Peters et al., 2004; Sanders et al., 2005). However, commercial mixtures of PBDEs can contain PBDDs and PBDFs that express significant AhR-mediated activities, such as CYP1A1 induction (Birnbaum et al., 2003; Hakk and Letcher, 2003). The expert panel had concerns about earlier results in the literature, indicating that PBDEs cause AhR-mediated effects because of the possible impurity issue similar to that described for the mono-ortho PCBs. In addition, it was also recognized that photochemical and combustion processes of PBDEs can also produce PBDDs and PBDFs. In conclusion, it was recommended that TEFs should not be assigned for PBDEs.
"Non-dioxin-like" AHR ligands and the TEF concept
The AhR can bind and be activated by a structurally diverse range of synthetic and naturally occurring chemicals (Denison and Heath-Pagliuso, 1998; Heath-Pagliuso et al., 1998; Jeuken et al., 2003; Nagy et al., 2002). These chemicals are widely distributed in dietary vegetables, fruits, teas, and dietary herbal supplements sometimes at relatively high concentrations (Amakura et al., 2002; Berhow et al., 1998; Formica and Regelson, 1995; Herzog et al., 1993a,b; Jeuken et al., 2003). The ability of metabolically labile phytochemicals to induce or inhibit induction of CYP1A1-dependent activities by 2,3,7,8- TCDD in cell culture model systems have been reported by numerous laboratories (Amakura et al., 2002; Jeuken et al., 2003; Williams et al., 1993; Zhang et al., 2003). However, the majority of toxicity studies demonstrated that these naturally occurring AhR agonists fail to produce AhR-dependent toxicity (Leibelt et al., 2003; Pohjanvirta et al., 2002), although some developmental dioxin-like effects have been reported for indole-3-carbinol (I3C) (Wilker et al., 1996). In addition, naturally occurring AhR ligands, such as I3C and diindolymethane, have been reported to inhibit 2,3,7,8-TCDD–dependent in vivo induction of CYP1A1 and immunotoxicity (Chen et al., 1995, 1996).
The ability of some non-dioxin–like PCBs and PCDFs to inhibit 2,3,7,8-TCDD–induced CYP1A1 activity and immunotoxicity in C57BL/6J mice has also been reported (Bannister and Safe, 1987; Biegel et al., 1989; Chen and Bunce, 2004; Crofton et al., 2005; Davis and Safe, 1988; Loeffler and Peterson, 1999; Morrissey et al., 1992; Smialowicz et al., 1997), whereas other studies have shown synergistic effects on dioxin toxicity of non-dioxin–like compounds, e.g., thyroid hormones, porphyrins, reproductive toxicity, and immunotoxicity (Bannister and Safe, 1987; Birnbaum et al., 1986; Crofton et al., 2005; Loeffler and Peterson, 1999; van Birgelen et al., 1996b).
The above studies provide evidence that non-dioxin–like compounds that are weak AhR agonists can modulate the overall toxic potency of 2,3,7,8-TCDD and related compounds. If occurring under natural background situations, these interactions might impact the magnitude and overall toxic effects produced by a defined amount of TEQ (i.e., from intake or present in the body) but not impact the determination of individual REP or TEF values for dioxin-like chemicals. The potential impact of these non-dioxin–like natural compounds on the risk of toxicity posed by exposure to a particular level of TEQs should be further investigated.
Additivity, an important prerequisite of the TEF concept was found to be consistent with results from recent in vivo mixture studies (Fattore et al., 2000; Gao et al., 1999;Hammet al., 2003; Walker et al., 2005). These studies showed thatWHO1998 TEF values predicted mixture toxicity within a factor two or less. Such accuracy is almost surprising in view of the fact that TEFs are derived from a range of REPs using different biological models or endpoints and are considered estimates with an order of magnitude uncertainty (Van den Berg et al., 1998).
The expert panel recognized that there are studies providing evidence that non-dioxin–like AhR agonists and antagonists are able to increase or decrease the toxicity of 2,3,7,8-TCDD and related compounds. Accordingly, their possible effect on the overall accuracy of the estimated magnitude of the TEQ needs to be investigated further, but it does not impact the experimental determination of individual REPs or TEFs.
For this TEF reevaluation process, the expert panel made extensive use of the refined TEF database that was recently published by Haws et al. (2006). Decisions about a TEF value were based on a combination of unweighted REP distributions, expert judgment, and point estimates. The use of solely unweighted REP distributions to set a TEF value was rejected because a specific percentile would have to be used as a cutoff, which could equally well be considered as a point estimate. However, such a percentile would have a lower biological or toxicological relevance than that obtained by expert judgment.
Previous TEFs were assigned in increments of 0.01, 0.05, 0.1, etc., but for this reevaluation, it was decided to use half order of magnitude increments on a logarithmic scale at 0.03, 0.1, 0.3, etc. This should be more useful in describing, with statistical methods, the uncertainty of TEFs in the future. In Table 1, the WHO 1998 and 2005 TEF values are summarized.
Figure 4 gives some indication of the quantitative impact of the 2005 changes on WHO TEF values in some selected biotic samples. The changes are shown as the ratio between the 2005 and 1998 WHO TEF values. In general, it can be concluded that the changes in 2005 values have a limited impact on the total TEQ of these samples with an overall decrease in TEQ ranging between 10 and 25%. The exception being the chicken where the decrease of the TEF for 2,3,4,7,8-PeCDF (from 0.5 to 0.3) and of lower TEFs for the mono-ortho PCBs resulted in an almost 50% decrease of total TEQ. In view of this average impact of 10–25%, it should be realized that many duplicate GC-MS analyses for these compounds also have an uncertainty that can fall in the range of 10–25%.
Several groups of compounds were identified for possible future inclusion in the TEF/TEQ concept. Based on mechanistic considerations, PCB 37, PBDDs, PBDFs, PXCDDs, PXCDFs, PCNs, PBNs, and PBBs undoubtedly belong in the TEF concept. However, for most, if not all, of these compounds there is a distinct lack of human exposure data. Therefore, preliminary exposure assessments should be done in the near future to indicate if these compounds are relevant for humans with respect to TEQ dietary intake. In addition, HCB could be a possible candidate for inclusion in the TEF/TEQ concept but only if it is unequivocally shown that impurities have not been the cause of earlier dioxin-like effects observed in experimental models. With respect to PBDEs, it was concluded that there is no reason for their inclusion in the TEF/TEQ concept.
Concern is expressed about the application of the TEF/TEQ approach to abiotic environmental matrices, such as soil, sediment, etc. The present TEF scheme (see Table 1) and TEQ methodology are primarily meant for estimating exposure via dietary intake situations because present TEFs are based largely on oral uptake studies often through diet. Application of these ‘‘intake or ingestion’’ TEFs for calculating the TEQ in abiotic environmental matrices has limited toxicological relevance and use for risk assessment, unless the aspect of reduced bioavailability and environmental fate and transport of the various dioxin-like compounds are taken into account. If human risk assessment is done for abiotic matrices, it is recommended that congener-specific equations be used throughout the whole model, instead of using a total TEQ basis, because fate and transport properties differ widely between congeners.
A number of future approaches to determine alternative or additional TEFs were identified. The use of a probabilistic methodology to determine TEFs has the advantage that it better describes the level of uncertainty in a TEF. The disadvantage could be that this approach lumps data together and gives similar weight to all studies, a problem that can only partly be avoided by separating in vitro from in vivo REPs. In addition, the sole use of a probabilistic approach includes other decision points, e.g., establishing a range of values from which one or more TEF values could be selected for risk assessment. Clearly, such an approach might cause problems for regulatory authorities and international harmonization of TEFs. Furthermore, choosing a specific percentile (e.g., 50th, 75th, or 95th) would, in fact, not be far different from using a point estimate.
The use of the present TEF values with body burden matrices as blood and adipose tissue have certain caveats from a risk assessment point of view as they were determined from intake situations. There is emerging experimental evidence which suggests that some REPs may differ when based on administered dose versus tissue concentration. The development and use of systemic TEFs and TEQ are recommended as an additional approach to the present TEF concept, but at present, there are insufficient data to develop these systemic TEFs.
- TEF concept
- Ah receptor ligands
- TEF concept in environmental health assessment
- TEF concept and uncertainty analysis
- Additivity of TEFs
- TEF concept and ecology
- TEF concept and Ah receptor
- Estimating TEF values
relative unit (compared with TCDD toxicity)
|Isomer groups||Toxicity factor relative to 2,3,7,8-T4 CDD|
a OME, 1984. b Excluding 2,3,7,8-T4CDD.
|Congener||EPA/87 a||NATO/89 b||WHO/94 c||WHO/98 d||WHO/2005 e|
|PCBs; IUPAC #, Structure|
a U.S. EPA, 1987. b NATO/CCMS, 1989. c Alhlborg et al., 1994. d Van den Berg, 1998. e Van den Berg, 2006.
|Chemical||Number of in vivo endpoints||Range of REPs (mean±std)||Number of end points from subchronic studies||Range of REPs (mean±std)||TEF|
|1,2,3,7,8-PCDD||22||0.16-0.9 (0.5±0.22)||16||0.19-0.9 (0.53±0.24)||1|
|2,3,4,7,8-PCDF||40||0.018-4.0 (0.4±0.7)||20||0.018-0.6 (0.20±0.13)||0.5|
|PCB 126||62||0.0024-0.98 (0.20±0.20)||31||0.004-0.18 (0.13±0.13)||0.1|
- U.S.EPA (2003): Toxic Equivalency Factors (TEF) for Dioxin and Related Compounds. In: Exposure and Human Health Reassessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and Related Compounds. Part II: Health Assessment for 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and Related Compounds. Chapter 9. NAS Review Draft NCEA-I-0836. December 2003. www.epa.gov/ncea.
DISCLAIMER This document is a draft. It has not been formally released by the U.S. Environmental Protection Agency and should not at this stage be construed to represent Agency policy. Mention of trade names or commercial products does not constitute endorsement or recommendation for use.
For additional references, see Toxic equivalency factor references.