Additivity of TEFs
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9.4. TOTAL TEQ AND THE ADDITIVITY CONCEPT
The issue of the scientific defensibility of additivity in determining total TEQ has been raised since the onset of the use of TEFs. Arguments regarding this approach include the presence of competing agonists or antagonists in various complex mixtures from environmental sources, interactions based on non-dioxin-like activities (inhibition or synergy), and the fact that dose-response curves for various effects may not be parallel for all congeners assigned TEFs. Although comparative pharmacokinetics have also been raised as an issue, this has generally been accounted for by the heavier weight accorded to in vivo studies in the assignment of TEFs. Despite these concerns, empirical data support the use of the additivity concept, recognizing the imprecise nature of the TEFs per se. A substantial effort has been made to test the assumptions of additivity and the ability of the TEF methodology to predict the effects of mixtures of dioxin-like compounds. These efforts have focused on environmental, commercial, and laboratory-derived mixtures. In addition, endpoints examined ranged from biochemical alterations, such as enzyme induction, to toxic responses such as tumor promotion, teratogenicity, and immunotoxicity. A brief summary of some of the more important work is given and discussed in the following section.
The TEF methodology has been examined by testing mixtures of chemicals containing dioxins and sometimes other chemicals. These mixtures have either been combined and produced in the laboratory or were actual environmental samples. Researchers have also used different approaches in estimating the TCDD equivalents of the mixtures. Some researchers have determined the REP of the components of the mixture in the same system in which the mixture was tested and have used these REPs to estimate TCDD equivalents. These studies can provide insight into the validity of the assumption of additivity of the TEF methodology. Other researchers have used consensus TEF values to estimate the TCDD equivalents of the mixture. It is not clear if these studies can be considered true tests of the additivity assumption. The consensus TEF values have been described as conservative estimates of the relative potency of a chemical in order to protect humans and wildlife. If the consensus TEF values are conservative and protective, then they should overestimate the potency of mixtures tested in an experimental system. In essence, using the consensus TEF values should generally over predict the potency of a mixture (and therefore under predict the response) when compared to the equivalent concentrations of TCDD in an experimental system. In the following discussion of the studies examining the assumption of additivity, these differences in study design and their implications for interpretation of the data must be considered.
9.4.1. Examination of Laboratory Mixtures of PCDDs and PCDFs
Bock and colleagues evaluated the TEF methodology in several systems using both individual congeners as well as laboratory-derived mixtures (Lipp et al., 1992; Schrenk et al., 1991, 1994). REPs or toxic equivalents or "TEs" (as designated by the authors) were determined for 2,3,7,8-substituted PCDDs based on enzyme induction in human HepG2 cells, rat H4IIE cells, and primary rat hepatocytes. Three laboratory-defined mixtures (M1, M2, and M3) were prepared and then examined in these same cell culture systems. TCDD contributed between 6%-8% of the TEQs for M1 and M2, but was not present in M3. In M1, 1,2,3,4,6,7,8-HpCDD contributes approximately 60% of the TEQ, and 1,2,3,7,8-PCDD and 1,2,3,4,7,8-HxCDD contribute 10% each. In M2, 1,2,3,4,6,7,8-HpCDD contributes 45%, 1,2,3,7,8-PCDD and 1,2,3,4,7,8-HxCDD contribute 15% each; and TCDD, 1,2,3,6,7,8-HxCDD, and 1,2,3,7,8,9-HxCDD contribute less than 10% to the total TEQ. The TEQs in M3 are derived predominately from 1,2,3,4,7,8-HxCDD (50%); 1,2,3,4,7,8-HxCDD (20%); and 1,2,3,6,7,8-HxCDD (18%). These mixtures also contain up to 49 chlorinated dibenzo-p-dioxins. The TCDD equivalents of the mixtures were determined on the basis of the assumption of additivity using the TEF methodology and the laboratory derived REPs or TEs as well as experimentally by comparing the EC50s of the mixtures with that of TCDD. According to the authors, in all three systems the data demonstrated that the components of the mixture act in an additive manner (Lipp, 1991; Schrenk et al., 1991). For example, in the human HepG2 cells the EC50 of a mixture of 49 different PCDDs was determined experimentally at 0.034 pg TEQ/plate, compared to the calculated or predicted EC50 of 0.028 pg TEQ/plate. Interestingly, the TEF methodology accurately predicted the effects of M3, a mixture containing predominately OCDD, some heptaCDDs and hexaCDDs, and no pentaCDDs or TCDD (Schrenck et al., 1991).
Bock and colleagues also tested a mixture of 49 PCDDs in a rat liver tumor promotion study. The mixture, designated as M2, was the same mixture used in the cell culture studies described above and TCDD contributed approximatley 8% of the TEQs of this mixture. In theses studies, rats received an estimated 2-200 ng TCDD/kg/d or 200-20,000 ng mixture/kg/d. The doses of the mixture were equivalent to the TCDD doses using a TE of the mixture of 0.01 based on enzyme induction in rat hepatocytes (Schrenk et al., 1991). A comparison of the relative potency of the mixture was based on liver concentrations of the chemicals followed by TEQ calculations using the I-TEFs (NATO/CCMS, 1988). According to the authors, in the low-dose region (2-20 ng TCDD/kg/d) the I-TEFs accurately predict the enzyme-inducing activity of the mixture but tend to overestimate the potency of the mixture at the higher doses (20-200 ng/kg/d). Also, according to the authors, the I-TEFs provide a rough estimate of the tumor-promoting potency of the mixture but overestimate the mixture’s potency . However, the authors did not quantify or qualify the magnitude of the overestimation.
In the studies by Schrenk and colleagues, the TEQs were based on tissue dose, not administered dose. Recent studies by DeVito et al. (1997b, 2000) indicate that the REP for dioxin-like compounds can differ when determined based on administered or tissue dose. The higher chlorinated dioxins tend to accumulate in hepatic tissue to a greater extent than does TCDD, and their REPs tend to decrease when estimated based on tissue dose (DeVito et al., 1997b, 2000). Because the I-TEFs are based on an administered dose, they may not predict the response when the TEQ dose is expressed as liver concentration. If the TEQ dose in the data by Schrenk et al. (1994) is compared on an administered dose, then the dose-response relationship for increases in relative volume of preneoplastic ATPase-deficient hepatic foci (% of liver) are comparable between TCDD and the mixture, indicating that additive TEFs provided an approximation of the tumor-promoting ability of a complex mixture of PCDDs (Schrenck et al., 1994). In addition, because TCDD contributed less than 10% of the total TEQ in these mixtures, these data indicate that the assumption of additivity reasonably predicts the response of complex mixtures of dioxins.
In responsive mouse strains, induction of cleft palate and hydronephrosis by TCDD occurs at doses between 3 and 90 :g TCDD/kg (Nagao et al., 1993; Weber et al., 1985; Birnbaum et al., 1985, 1987, 1991). Several groups have examined the assumption of additivity using teratogenic effects of dioxins as an endpoint. Birnbaum and colleagues examined TEF methodology using mouse teratogenicity as an endpoint (Weber et al., 1985; Birnbaum et al., 1985, 1987, 1991). REPs were derived for 2,3,7,8-TCDF, 2,3,4,7,8-PeCDF, 1,2,3,7,8-PeCDF, and 1,2,3,4,7,8-HxCDF (Weber et al., 1984, 1985; Birnbaum et al., 1987). Analysis of the dose-response for these chemicals, based on administered dose, demonstrated parallel slopes. According to the authors, dose-response analysis of two mixtures containing either TCDD and 2,3,7,8-TCDF or 2,3,4,7,8-PeCDF and 1,2,3,4,7,8-HxCDF demonstrated strict additivity (Birnbaum et al., 1987; Weber et al., 1985).
Nagao et al. (1993) also examined the TEF methodology using teratogenicity in mice as an endpoint. Mice were exposed to a single dose of TCDD (5-90 :g/kg) or a mixture of PCDDs, or one of two different mixtures of PCDFs. The mixtures contained no detectable TCDD. The I-TEFs were used to determine the TEQ of the mixtures. According to the authors, the I-TEFs predicted the potency of the PCDD mixture, and the dose-response relationship was consistent with the assumption of additivity. The I-TEFs overestimated the potency of the PCDF mixtures by two- or fourfold. All three mixtures contained significant concentrations of non 2,3,7,8-chloro-substituted PCDDs and PCDFs in addition to the dioxin-like compounds present. In the studies by Birnbaum and colleagues (Weber et al., 1985; Birnbaum et al., 1985, 1987, 1991) and Nagao et al. (1993) examining the assumption of additivity using teratogenicity as an endpoint, the TEF methodology proves useful in estimating the effects of these mixtures.
Rozman and colleagues have examined the assumption of additivity of PCDDs in both acute and subchronic studies. In acute studies, TCDD (20-60 :g/kg), 1,2,3,7,8-PCDD (100-300 :g/kg), 1,2,3,4,7,8-HxCDD (700-1,400 :g/kg), and 1,2,3,4,6,7,8-HpCDD (3,000-8,000 :g/kg) were administered to male rats, and REP values were determined for lethality. A mixture of all four chemicals at equally potent concentrations was then prepared and dose-response studies were performed with the mixture at doses that would produce 20%, 50%, and 80% mortality. The mixture studies demonstrated strict additivity of these four chemicals for biochemical and toxicological effects (Stahl et al., 1992; Weber et al, 1992a,b). Following the acute studies, Viluksela et al. (1998a,b) prepared a mixture of these chemicals and estimated the TEQ based on the REPs from the acute studies. A loading/maintenance dose regimen was used for 90 days and the animals were followed for an additional 90 days. According to the authors, the assumption of additivity predicted the response of the mixture for lethality, wasting, hemorrhage, and anemia, as well as numerous biochemical alterations such as induction of hepatic EROD activity and decreases in hepatic phosphenolpyruvate carboxykinase and hepatic tryptophan 2,3-dioxygenase (Viluksela et al., 1997; 1998). Increases in serum tryptophan concentrations and decreases in serum thyroxine concentrations were also predicted by the TEF methodology (Viluksila et al., 1998a).
Rozman and colleagues followed up these initial studies by examining the assumption of additivity of the effects of PCDDs as endocrine disruptors (Gao et al., 1999). Ovulation is a complex physiological phenomenon that requires the coordinated interaction of numerous endocrine hormones. In a rat model, ovulation can be inhibited by TCDD at doses between 2 to 32 µg/kg (Gao et al., 1999). Dose-response analysis of TCCD, 1,2,3,7,8-PeCDD, and 1,2,3,4,7,8-HxCDD demonstrate that the slopes are parallel and the REPs are 0.2 and 0.04, respectively. According to the authors, the dose response for a mixture of these chemicals, in which the components were at equally potent concentrations, further demonstrated the response additivity of mixtures of PCDDs and the predictive ability of the TEF methodology (Gao et al., 1999).
The research on the interactions between mixtures of PCDDs and PCDFs has taken two approaches. The first is to derive REP values in the same system in which the mixtures shall be tested. These studies confirm that the assumption of additivity can predict the response of mixtures of PCDDs and PCDFs. A second approach is to use the I-TEFs to assess the potency of a mixture. These studies tend to indicate that the I-TEFs overestimate the potency of a mixture by factors of two to four. Recently, the WHO TEFs have been described as "order of magnitude" estimates of the potency of dioxin-like compounds. However, the studies using consensus TEFs demonstrate that for mixtures of PCDDs and PCDFs, the TEF methodology will predict within a half-order of magnitude or less (Schrenck et al., 1994; Nagao et al., 1993). In either case, the TEF methodology accurately predicts the responses of experimentally defined mixtures of PCDDs and PCDFs. Furthermore, several of these studies described the effects of mixtures containing either no TCDD or with TCDD contributing less than 10% of the TEQ in the presence of significant concentrations of non-2378- CDDs and CDFs. These studies strongly support the use of the TEF methodology.
9.4.2. Examination of Commercial or Laboratory-Derived Mixtures of PCDDs, PCDFs, and PCBs
Commercial mixtures of PCBs elicit a broad spectrum of biological and toxicological responses in both experimental animals and humans. Some of the observed effects resemble those induced by dioxin and furans (enzyme induction, immunotoxicity, teratogenicity, endocrine alterations, etc.). Attempts to expand the TEF approach to risk assessment of PCBs have investigated the ability of both commercial PCBs and individual congeners, selected on the basis of structure-activity relationships, to induce dioxin-like effects and to interact with TCDD. One of the first studies to examine the interactions of individual PCB congeners with TCDD used mouse teratogencity as an endpoint (Birnbaum et al., 1985, 1987). A mono-ortho PCB (2,3,4,5,3',4'-HxPCB or PCB 156) at doses of 20 mg/kg or higher (Birnbaum, 1991) induced hydronephrosis and cleft palate in mice. When mice were co-exposed to PCB 156 and 3.0 :g TCDD/kg the interactions resulted in strict additivity.
The interaction of TCDD with dioxin-like PCBs has been examined by van Birgelen et al. (1994a,b) in subchronic rat feeding studies. Concentrations of PCB 126 in the diet between 7 and 180 ppb induced several dioxin-like effects, including CYP1A1 induction, thymic atrophy, liver enlargement, and decreases in hepatic retinol concentrations, body weight gains, and plasma thyroxine concentrations. The REP for PCB 126 was estimated by the authors at between 0.01 and 0.1 (van Birgelen et al., 1994a). Co-exposure to PCB 126 and TCDD (0.4 or 5.0 ppb) in the diet demonstrated additivity for all responses except induction of CYP1A2 and decreases in hepatic retinol, where antagonism occurred at the highest doses of PCB 126 and TCDD tested. These nonadditive interactions were not observed at more environmentally relevant exposures, according to the author. In a similar study design, PCB 156 also induced dioxin-like effects with a REP estimated between 0.00004 and 0.001 (van Birgelen et al., 1994b). Similar to the interactions between PCB 126 and TCDD, additive interactions were observed in animals receiving mixtures of PCB 156 and TCDD in the low-dose region for all responses examined. However, at the highest exposures of PCB 156 and TCDD, the authors reported slight antagonistic interactions for decreases in hepatic retinol (van Birgelen et al., 1994b). For both PCB 126 and PCB 156, antagonistic interactions were observed with TCDD only at exposures that produced maximal CYP1A1 induction. The authors concluded that the antagonistic interactions are unlikely to occur at relevant human exposures.
In a series of studies examining the TEF methodology, TCDD (1.5-150 ng/kg/d), 1,2,3,7,8-PeCDD; 2,3,7,8-TCDF; 1,2,3,7,8-PeCDF; 2,3,4,7,8-PeCDF; OCDF; the coplanar PCBs 77, 126, and 169; and the mono-ortho substituted PCBs 105, 118, and 156 were administered to mice 5 days/week for 13 weeks. REPs were determined for EROD induction, a marker for CYP1A1, in liver, lung, and skin; ACOH activity, a marker for CYP1A2, in liver; and hepatic porphyrins (DeVito et al., 1997a; 2000; van Birgelen et al., 1996c). These data demonstrate that the dose-response curves for the PCDDs and PCDFs were parallel (DeVito et al., 1997a). Dose-response curves for some of the enzyme induction data for the individual PCBs displayed evidence of non-parallelism in the high-dose region (DeVito et al., 2000). A laboratory-derived mixture of these chemicals with congener mass ratios resembling those in food was administered to mice and rats, and indicated that despite the evidence of non- parallelism for the PCBs at high doses, the assumption of additivity predicted the potency of the mixture for enzyme induction, immunotoxicity, and decreases in hepatic retinoids (Birnbaum and DeVito, 1995; van Birgelen et al., 1996; 1997; DeVito et al., 1997; Smialowicz et al., 1996). In addition, the REPs estimated in mice also predicted the response of the mixture in rats for enzyme induction and decreases in hepatic retinyl palmitate concentrations (van Birgelen et al., 1997d; Ross et al., 1997; DeVito et al., 1997b). These studies indicate that not only do the REPs for enzyme induction in mice predict other responses, such as immunotoxicity and decreases in hepatic retinyl palmitate, they also can be used to predict responses of mixtures in another species.
The commercial PCB mixtures induce a variety of dioxin-like effects. Rats exposed to commercial Aroclors and observed for 2 weeks exhibited dose-dependent induction of hepatic CYP1A activity (EROD) but no thymic atrophy (Harris et al., 1993). Using REP values derived for EROD induction in rats, the TEF methodology provided good agreement with experimental estimates of the ED50 for enzyme induction. However, use of the conservative TEF values of Safe (1990) overestimated the potency of the Aroclor mixutres (Harris et al., 1993). In contrast, similar studies examining immunotoxicity as an endpoint demonstrate that both experimentally derived REP values and the conservative TEF values of Safe (1990) overestimate the potency of the Aroclor mixtures by a factor of 1.2 - 22 (Harper et al., 1995). These data demonstrate that there are nonadditive interactions between dioxin-like compounds and the non-dioxin-like PCBs and that these interactions are response specific and most likely are not due to AhR antagonism.
In in vitro systems, using H4IIe cells and rat hepatocytes, Schmitz et al. (1995, 1996) examined the assumption of additivity for individual congeners as well as commercial mixtures. After deriving REP values for enzyme induction, the authors concluded that a laboratory mixture of PCBs 77, 105, 118, 126, 156, and 169 demonstrated perfect additive behavior in these cell line systems (Schmitz et al., 1995). However, when the mixture was combined with a tenfold surplus of a mixture containing non-dioxin-like PCBs (PCB 28, 52, 101, 138, 153 and 180), the mixture demonstrated an approximate threefold higher TEQ than predicted. The authors concluded that a moderate synergistic interaction is responsible for the increased enzyme-inducing potency of the mixture containing dioxins and non-dioxin-like PCBs. Further studies by Schmitz et al. (1996) also demonstrated a slight synergistic deviation (less than threefold) from strict additivity when the calculated TEQ based on chemical analysis of Aroclor 1254 and Clophen A50 was compared to the CYP1A-induction TEQ derived in an established rat hepatoma cell line (H4IIE) (Schmitz et al., 1996).
Recently, Mayes et al. (1998) compared the carcinogenicity of Aroclor 1016, 1242, 1254 and 1260 in Sprague-Dawley rats. All four mixtures increased the incidence of hepatic tumors in female rats. The authors concluded that the female rats were more susceptible than the males to the hepatocarcinogenic effects of these mixtures. In the two-year bioassay of TCDD in Sprague-Dawley rats, the female rats were also more susceptible to the hepatocarcinogenic effects than the males (Kociba et al., 1978). Mayes and colleagues(1998) performed congener specific analysis of the Aroclor mixtures and calculated dioxin TEQ values for each of these mixtures. In order to compare the cancer induction potential of dioxin TEQ in PCB mixtures (Mayes et al. 1998) with that from TCDD (Kociba et al., 1978) in the same species of rat, the dose-response relationships are graphed and presented in figure 9-2. The dose-response relationship for hepatic tumors in female rats is similar between the Aroclor 1242, 1254, 1260 and TCDD dose regimen. This analysis demonstrates that the TEF methodology qualitatively and quantitatively predicts the response of a complex mixture of PCBs. This is particularly important because the mass concentration of dioxin equivalents in the mixture is approximately 100,000 times less than the non-dioxin-like PCBs present in these mixtures. These data strongly support the ability of the TEF methodology to estimate the carcinogenic potency of a complex mixture of PCBs even in the presence of significant concentrations of non-dioxin-like PCBs.
Researchers have evaluated the applicability of the TEF methodology to mixtures containing dioxin-like PCBs by examining the interactions of binary mixtures, laboratory-derived mixtures, or commercial mixtures of PCBs. The studies examining the binary mixtures or laboratory-derived mixtures have demonstrated that the assumption of additivity provides good estimates of the potency of a mixture of PCBs and other dioxin-like compounds. In contrast, studies using commercial mixtures of PCBs suggest that the assumption of additivity may be endpoint specific, and that both synergistic and antagonistic interactions may occur for some mixtures of dioxins and PCBs for certain endpoints. A more detailed examination of these issues follows in the section on nonadditive interactions with non-dioxin-like compounds.
9.4.3. Examination of Environmental Samples Containing PCDDs, PCDFs, and/or PCBs
One of the first tests of the TEF methodology examined soot from a transformer fire in Binghamton, NY (Eadon et al., 1986). Benzene extracts of soot from a PCB transformer fire which contained a complex mixture of PCDDs, PCDFs, PCBs, and polychlorinated biphenylenes were administered to guinea pigs as single oral doses, and LD50 values were compared to TCDD. Relative potency values for the PCDDs and PCDFs based on guinea pig LD50 values were used to estimate the TCDD equivalents of the mixture. Eadon and co-workers exposed guinea pigs to either TCDD alone or the soot and determined their LD50s. With these relative potency values, the soot extract had a TCDD equivalent concentration of 22 ppm. Comparison of the LD50s for TCDD and the soot led to a TCDD equivalent of 58 ppm for the mixture. Other endpoints examined included alterations in thymus weight, body weight, serum enzymes, and hepatotoxicity. Experimentally the TCDD equivalents of the soot varied from 2 to 58 ppm. The authors concluded that because the benzene extract of the soot contained hundreds of chemicals including PCDDs, PCDFs, and PCBs, the difference between the calculated TEQ of 22 ppm and the experimentally derived TEQs between 2 and 58 seems minimal. (Note: the initial analytical TEQ value of soot [22 ppm] was calculated on the basis of guinea pig LD50 values of the respective components; using the current recommended TEF scheme [van den Berg et al., 1998], the "calculated" TCDD TEQ would be approximately 17 ppm.)
Shortly after the studies on the Binghamton transformer fire soot, investigators applied the TEF methodology to the leachate from Love Canal, NY. The organic phase of the leachate consisted of more than 100 different organic compounds including PCDDs and PCDFs. The leachate did not contain PCBs or PAHs. The authors estimated the TEQ of the mixture on the basis of REP values for teratogenicity (cleft palate and hydronephrosis in mice) for the PCDDs and PCDFs present in the leachate. The authors state that the leachate contained the equivalent of 3 :g TCDD/g and that more than 95% of the TEQ was contributed by TCDD. There were two other PCDFs present in the leachate, and their contribution to the total TEQ was approximately 5% (Silkworth et al., 1989). When the TEQ of the mixture was based on dose-response analysis of the mixture compared to TCDD, the leachate was estimated to contain between 6.6 and 10.5 :g TCDD/g (Silkworth et al., 1989). The authors concluded there was a good agreement between the experimental TCDD equivalents (6.6-10.5 :g TCDD/g) and the analytical TEQs (3 :g TCDD/g). In addition, these studies illustrate that the non-AhR components of the leachate did not interfere with receptor-mediated teratogenicity (Silkworth et al., 1989). Additional investigations have shown that the same complex mixture of non-AhR agonists slightly potentiated TCDD-induced thymic atrophy and immunosuppression (plaque-forming cells/spleen response) while decreasing the hepatic CYP1A-inducing ability of the TCDD component (Silkworth et al., 1993).
The assumption of additivity was also examined using a PCDD/PCDF mixture extracted from fly ash from a municipal waste incinerator (Suter-Hofmann and Schlatter, 1989). As a purification step, rabbits were fed organic extracts from the fly ash. After 10 days the livers were removed and analyzed for PCDDs and PCDFs. The rabbit livers contained predominately 2,3,7,8-substituted PCDDs/PCDFs. Based on the chemical analysis of the liver, lyophilized and pulverized liver was added to the standard rat diet. This diet was fed to rats for 13 weeks and body weights and terminal thymus weights were recorded. The authors concluded that the mixture of PCDDs and PCDFs produced equivalent toxicities as TCDD, and the assumption of additivity was confirmed.
9.4.4. Nonadditive Interactions With Non-Dioxin-Like Compounds
For a number of toxicological responses, there appears to be evidence for nonadditive interactions in defined dose ranges by both commercial Aroclors and major congeners with little if any AhR agonist activity (i.e., PCB 153). Both commercial Aroclors and a PCB mixture comprised of major congeners found in human breast milk were shown to antagonize the immunotoxic effects of TCDD in mice (Biegel et al., 1989; Davis and Safe, 1989; Harper et al., 1995). When immunotoxicity-derived TEF values for a variety of PCB congeners were used in an additive manner to estimate TCDD TEQs for commercial Aroclors, in comparison to the experimental TEQs, they were approximately predictive for Aroclor 1254 and 1260 (Harper et al., 1995). However, the TEF approach tended to overestimate the immunotoxicity of Aroclors 1242 and 1248, suggesting some antagonism.
Typical responses to TCDD exposure in rodents include CYP1 enzyme induction and thymic atrophy. Rats consuming a diet containing 5 ppb TCDD for 13 weeks exhibited a 33-fold increase in hepatic CYP1A activity (EROD) and a greater than 50% reduction in relative thymus weight. Addition of PCB 153 to the diet at concentrations up to 100 ppm had no significant effect on either response (van der Kolk et al., 1992). Mice dosed simultaneously with TCDD and up to a 106-fold molar excess of PCB 153 (1 nmol/kg vs. 1 mmol/kg) exhibited no significant dose-dependent alteration in hepatic CYP1A1/A2 protein compared to the TCDD dose group alone (De Jongh et al., 1995). There was, however, an approximate twofold increase in hepatic EROD activity in the highest combined PCB 153:TCDD dose group. Subsequent tissue analysis revealed that the increase in EROD activity appeared related to PCB 153 increasing hepatic TCDD concentrations. The same PCB congener at high doses (358 mg/kg) is able to almost completely inhibit TCDD-induced suppression of the plaque-forming cell (PFC) response toward sheep red blood cells in male C57BL/6J mice (Biegel et al., 1989; Smialowicz et al., 1997). However, as PCB 153 displays negligible AhR binding affinity, the exact mechanism(s) behind these interactions is unknown. Recently, it has been shown that PCB 153 at high doses (greater than 100 mg/kg) actually enhances the PFC response in female B6C3F1 mice, thereby raising the "control" set point. When combined doses of TCDD and PCB 153 are then compared to the elevated PCB 153 response, an apparent block of the immunosuppressive effect of TCDD is observed (Smialowicz et al., 1997). The relevance of this functional antagonism is uncertain, as the doses required to inhibit the TCDD-like effects are at least 100 mg/kg of PCB 153. These doses of PCB 153 seem unlikely to occur in human populations except under extreme conditions.
Commercial PCBs and various PCB congeners have been shown to potentiate or antagonize the teratogenicity of TCDD depending upon the dose ranges and response examined (Biegel et al., 1989; Morrissey et al., 1992). Treatment of developing chicken embryos with TCDD and dioxin-like PCBs induces a characteristic series of responses, including embryo lethality and a variety of embryo malformations/deformities. Combined exposure of chicken embryos to both PCB 126 and PCB 153 (2 :g/kg and 25-50 mg/kg, respectively) resulted in protection from PCB 126-induced embryo malformations, edema, and liver lesions, but not mortality (Zhao et al., 1997). In mice, doses of 125 mg PCB 153/kg or higher inhibit the induction of cleft palate by TCDD (Biegel et al., 1989; Morrissey et al., 1992). The induction of hydronephrosis by TCDD was slightly antagonized by PCB 153, but only at doses of 500 mg/kg or higher. Once again, the environmental relevance of exposures of 100 mg/kg of PCB 153 or higher remains quite speculative, and nonadditive interactions are not expected at environmental exposures.
Nonadditive interactions have also been observed in rodents exposed to both TCDD and mixtures of various PCB congeners for hepatic porphyrin accumulation and alterations in circulating levels of thyroid hormones. A strong synergistic response was seen with hepatic porphyrin accumulation in female rats following the combined dietary exposure to TCDD and PCB 153 (van Birgelen, 1996a). The mechanism accounting for the interaction was thought to be a combination of both AhR-dependent (CYP1A2 induction) and AhR-independent (*-aminolevulinic acid synthetase [ALAS] induction) events. Additionally, subchronic exposure of mice to a mixture of PCDDs, PCDFs, and dioxin-like PCBs in a ratio derived from common foods also resulted in a highly synergistic response, when compared to an equivalent dose of TCDD alone, for both hepatic porphyrin accumulation and urinary porphyrin excretion (van Birgelen et al., 1996b). PCB 153, although not porphyrinogenic alone, when added to the mixture further enhanced the synergistic response of hepatic porphyrin accumulation. Non-AhR-mediated induction of ALAS activity by both the dioxin-like mono ortho-substituted PCBs in the mixture and by PCB 153 was hypothesized to partially explain the synergism.
Decreases in thyroid hormone levels have been observed in both experimental animals and humans following exposure to both dioxin-like and non-dioxin-like compounds (Nagayama et al., 1998; Koopman-Esseboom et al., 1997). It is currently thought that multiple mechanisms, including induction of specific isozymes of hepatic UDP-glucoronyl transferase (UDPGT) and binding to thyroid hormone transport proteins (thyroid binding globulin, transthryetin) could be involved. Exposure of female rats to a food-related mixture of PCDDs, PCDFs, and dioxin-like PCBs for 90 days resulted in an approximately 85% decrease in decrease in plasma levels of thyroxine. In contrast, the TCDD equivalent dose produced no effect on serum thyroxine (van Birgelen et al., 1997). Increased induction of several isoforms of UDPGT by the HAH mixture as compared to TCDD was thought to only partially explain the observed response with thyroxine levels.
Several studies examining the interactions of dioxins and non-dioxins for rat liver tumor promotion and additive and nonadditive interactions have been reported. Synergistic interactions for tumor promotion have been observed for combinations of PCB 77 and PCB 52 (2,2',5,5'-tetrachlorbiphenyl) in rat liver (Sargent et al., 1992). Bager et al. (1995) reported greater than additive interactions of PCBs 126 and 153 in a rat liver tumor promotion model.
The assumption of additivity was examined in a laboratory-derived mixture of PCDDs, PCDFs, and PCBs in a rat liver tumor promotion model (van der Plas et al., 1999). The mixture contained TCDD, 1,2,3,7,8-PeCDD, 2,3,4,7,8-PeCDF, and PCBs 126, 118, and 156. The composition of the mixture was based on concentrations of these chemicals in Baltic herring. PCB 126 and 1,2,3,7,8-PeCDD accounted for 65% of the TEQ in the mixture and TCDD accounted for approximately 6.6%. Both TCDD and the TEQ mixture increased mean foci volume and the volume fraction of foci in the liver. However, the response was statistically significantly greater in the TCDD treated animals by approximately 2-fold. While the TEQ mixture did not produce the exact same response level as TCDD, it is difficult to quantify the overestimation of the TEF methodology in this study since only a single dose level was examined. The authors also did a dose-response study with the mixture. However, they added PCB 153 to the mixture used for the dose response study. The concentration of PCB 153 was 20,000 times the concentration of TCDD in these mixtures. Dose levels of 0.5, 1, and 2 ug TEQ/kg/week were administered to the animals. The presence of PCB 153 did not alter the effects of the 1 ug TEQ/kg/week dose since there was no statistical difference between the response of animals to the TEQ mixture with or without PCB 153. The highest dose examined, 2 ug TEQ/kg/week produced an effect that was statistically equivalent to the animals treated with TCDD alone. Van der Plas et al (1999) also determined the concentration of chemicals in the liver at the termination of the study. Their data suggest that the lower response level of the mixture is due to pharmacokinetic interactions. Animals administered 1 ug TEQ/kg/week had approximately one third of the liver TEQ concentrations as the animals treated with TCDD. Animals treated with 2 ug/kg/week had equivalent TEQ concentrations in the liver and also had similar responses as animal treated with 1 ug TCDD/kg/week. Van der Plas and colleagues concluded that the TEF methodology predicted the tumor-promoting potency of the mixture quite well, within a factor of two, but pharmacokinetic interactions between dioxins may alter the accuracy of the methodology (van der Plas et al., 1999).
In another study, van der Plas and colleagues (2000) examined the interactions of coplanar and non-coplanar components of Aroclor 1260 in a tumor promotion study. In these studies, Aroclor 1260 was separated into planar (0-1 ortho chlorines) and non-planar (2-4 ortho chlorines) components. Rats were then exposed to either 1 ug TCDD/kg/week, 1 mg 0-1ortho/kg/week, 9 mg 2-4 ortho/kg/week, 10 mg 0-4 ortho/kg/week or 10 mg aroclor 1260/kg/week. Mean foci volume and the volume fraction of the liver occupied by foci increased in animals treated with either TCDD, the 2-4 ortho mixture, the 0-4 ortho mixture and aroclor 1260. The 0-1 ortho mixture did not alter foci development compared to the control animals. Van der Plas et al (2000) concluded that the results from their study indicate that 80% of the carcinogenicity of Aroclor 1260 is due to the non-dioxin congeners in the mixture.
In the study described above, Van der Plas et al (2000) used the CALUX assay to determine the TEQ of the different mixtures. The lot of Aroclor 1260 used in this study had very low TEQs based on the CALUX assay. For example, 10 mg Aroclor 1260/kg/week was equivalent to 0.0012 ug TEQ/kg/week or approximately 0.12 ppm TEQ. In addition, the 1 mg 0-1 ortho/kg/week dose is equivalent to 0.09 ng TEQ/kg/d. In contrast, the lot of Aroclor 1260 used by Mayes et al (1998) had 7.2 ppm TEQ concentrations using the WHO TEF values and dose levels examined ranged from 10-42 ng TEQ/kg/d. The lot of Aroclor 1260 used by Mayes et al (1998) has approximately 60 times more TEQs than the lot used by van der Plas et al (2000). In the Mayes et al (1998) studies the TEF methodology accurately predicts the carcinogenic response of the mixture. The differences in the van der Plas et al (2000) and the Mayes et al (1998) studies may be due to the different lots of Aroclor 1260 used by these two groups.
The interactions of dioxins with non-dioxin-like compounds results in additive and nonadditive responses. The antagonistic interactions, while endpoint specific, appear to occur at dose levels that greatly exceed most human exposures and should not affect the overall use of the TEF methodology. One of the difficulties in addressing the nonadditive interactions is understanding the mechanism behind these interactions. For the greater than additive interactions for induction of porphyria and decreases in serum thyroxine, there are hypotheses that may explain these effects. The mechanism of the antagonistic interactions of non-dioxin-like PCBs and TCDD on immunotoxicity and teratogenicity in mice is uncertain. For other responses, such as developmental reproductive toxicity, the interactions of PCDDs, PCDFs, and PCBs have not been examined. In addition, it has also been suggested that antagonism of Ah receptor-mediated events may be species specific. For example, addition of PCB 52, a congener commonly found in biotic samples, inhibited the TCDD-induced expression of a reporter gene under the regulatory control of the Ah receptor in mouse and rat cells, but not in guinea pig or human hepatoma cells (Aarts et al., 1995). Our limited understanding of the interactions between dioxins and non-dioxins for a variety of responses requires further research before their impact on the TEF methodology can be fully understood.
9.4.6. Toxic Equivalency Functions
The TEF methodology has been described as an "interim" methodology. Since this interim method has been applied, there have been few proposed alternatives. One recent proposal suggests that the TEF value be replaced by a toxic equivalency function (Putzrath, 1997). It has been proposed that the REPs for PCDDs/PCDFs are better described by a function as compared to a factor or single-point estimate (Putzrath, 1996). The use of a factor to describe the relative potency of a chemical implies that its relative potency is independent of dose. Putzrath (1997) suggests that data exists which indicates that the REPs are dose dependent and that the REPs must be described as a function of dose. Recent studies have examined this possibility for a series of PCDDs/PCDFs and PCBs (DeVito et al., 1997; DeVito et al., 2000). For the PCDDs/PCDFs, the data indicate that the REPs estimated from enzyme induction data in mice are best described by a factor and not a function. For some of the PCBs examined, a function fit better, but the change in the REP was within a factor of two to five for most of the four enzymatic responses examined (DeVito et al., 2000). In addition, the dose dependency was observed only at the high-dose and not in the low-dose region (DeVito et al., 2000).
Even though these studies suggest that a TE function may be useful, there are numerous difficulties in applying this method. If the REPs are really functions and not factors, there must be a mechanistic basis for these differences, and these mechanisms would most likely be response specific and perhaps species specific. This would then require that for all critical responses, every chemical considered in the TEF methodology would have to be examined. Once again, it is highly unlikely that 2-year bioassays and multigenerational studies will be performed on all the TEF congeners in the foreseeable future. The use of a TEF function requires extensive data sets that are not available and are unlikely to be collected.
9.4.7. Species and Endpoint Specific TEFs
It is often suggested that species and endpoint TEFs may be required for the TEF concept to provide accurate estimates of risk. In fact, the WHO does have class specific TEFs based on fish, birds and mammals (van den Berg et al., 1998). The most notable differences are the lack of effect of some mono-ortho PCBs in fish (van den Berg et al., 1998). Hahn and colleagues recently examined the influence of affinity and intrinsic activity on the relative potency of PCBs in PLHC-1 cells (Hestermann et al., 2000). Using this cell line derived from fish, Hahn and colleagues demonstrated clear differences in the response of these cells to mono-ortho PCBs. The insensitivity of these fish cells to the mono-ortho PCBs is due to both reduced affinity and reduced intrinsic efficacy. Using information on affinity and intrinsic efficacy allowed for better predictions of mixtures of these chemicals than did the application of the TEF methodology (Hestermann et al., 2000). Future studies examining species differences applying the approach of Herstermann et al., (2000) may provide insight into species specific TEFs as well as alternative approaches to the TEF methodology.
There are numerous examples of endpoint specific relative potencies for receptor mediated pharmacological agents, such as the antiestrogen, tamoxifen. It is reasonable to assume that the Ah receptor and its ligands would be no different from these other receptor systems. Examination of the WHO data base suggests that even for the chemicals with the largest data sets this question cannot be adequately addressed (See section 9.2.5). Endpoint specific TEFs would require a much more complete data set than is available at this time. In addition, these studies would have to be designed to test the hypothesis that the REPs are equivalent across endpoints. This requires controlling the species and dosing regimen employed as well as other factors. One of the reasons the TEF methodology was developed was because limited toxicity data was available for the other dioxin-like chemicals and it was unlikely that all relevant chemicals would be tested for all responses in all species, including humans. For example, it is extremely unlikely that 2-year bioassays for carcinogenesis or multi-generational studies will be performed on all chemicals included in the TEF methodology. Even though there are significant data demonstrating that a number of chemicals produce dioxin-like toxic effects, clearly the data set is not complete. For this reason, WHO recommends revisiting the TEF values every 5 years.
- 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
- 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.