History of TEF concept

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A wide variety of polyhalogenated aromatic hydrocarbon (PHAH) compounds can be detected as complex mixtures in both abiotic and biotic samples. Because of PHAHs’ known global environmental distribution and their toxicity to experimental animals (DeVito et al., 1995; DeVito and Birnbaum, 1995; Grassman et al., 1998)(see Part II, Chapters 3-6), to wildlife (Giesy and Kannan, 1998; Ross, 2000), and to humans (IARC, 1997) (see also Part II, Chapter 7 ), hazard characterization and risk assessment activities have tended to focus on a subset of polychlorinated dibenzo-p-dioxin (PCDDs), polychlorinated dibenzofurans (PCDFs), and polychlorinated biphenyls (PCBs)(Figure 9-1). The subset of compounds, shown in Figure 9-1, are known as "dioxin-like" and have been assigned TEF values by WHO. In this chapter, the development of TEFs for these and other PHAHs is discussed.[1]

9.2.1. TEFs for PCDDs and PCDFs

In 1983, the Ontario Ministry of the Environment produced a Scientific Criteria Document for PCDDs and PCDFs which concluded, based on a review of available scientific information, that dioxin and dibenzofurans were structurally similar compounds that shared a common cellular mechanism of action (activation of the aryl hydrocarbon receptor or AhR) and induced comparable biological and toxic responses, and that the development of environmental standards for human health concerns should be based on a "toxic equivalency" approach with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) as the prototype (OME, 1984). The final recommendation divided all PCDD/PCDF congeners into their respective homologue groups and assigned to each group a toxicity factor relative to TCDD (Table 9-1). These numerical factors could then be applied to transform various concentrations of PCDDs and PCDFs into equivalent concentrations of 2,3,7,8-TCDD. Shortly thereafter, the first use of a TEF-like method was described by Eadon et al. (1986) as a means to estimate potential health risks associated with a PCB transformer fire in Binghamton, NY.

Following up on an initial risk assessment methodology designed to address the emission of dioxins and furans from waste incinerators, EPA also concluded that TEFs were the best available interim scientific policy for dealing with complex mixtures of these contaminants. With the mandate to develop active research programs that would address the limitations inherent to this risk management technique, the Agency recommended TEFs for specific congeners, rather than isomeric groups (Table 9-2; U.S. EPA, 1987). In an analogous fashion to OME's approach, concentrations of PCDDs and PCDFs would be analytically determined, the concentration of each congener would be multiplied by its respective TEF value, and all the products would be summed to give a single 2,3,7,8-TCDD equivalent. This approach has been described mathematically as:

Total Toxic Equivalency (TEQ) = Σ(n=1 to k) Cn * TEFn

Cn equals the concentration of the individual congener in the complex mixture under analysis. TEFs were determined by inspection of the available congener-specific data and an assignment of an "order of magnitude" estimate of relative toxicity when compared to 2,3,7,8-TCDD. In vitro AhR binding and in vitro and in vivo toxicity studies were considered in setting individual TEFs. Scientific judgment and expert opinion formed the basis for these TEF values. External review of the toxicity and pharmacokinetic data utilized by EPA in setting these TEFs supported the basic approach as a "reasonable estimate" of the relative toxicity of PCDDs and PCDFs (Olson et al., 1989).

A 3-year study conducted by the North Atlantic Treaty Organization Committee on the Challenges of Modern Society (NATO/CCMS) also concluded that the TEF approach was the best available interim measure for PCDD/PCDF risk assessment. On the basis of examination of the available data dealing with exposure, hazard assessment, and analytical methodologies related to dioxin and furans, an International Toxicity Equivalency Factor (I-TEF) scheme was presented (Table 9-2; NATO/CCMS, 1988). This review also concluded that "data strongly support the role of the AhR in mediating the biologic and toxic responses elicited by 2,3,7,8-TCDD and related PCDDs and PCDFs and provide the scientific basis for the development of TEFs for this class of compounds." Various refinements to previous efforts included selection of TEF values based more on in vivo toxicities, assigning TEF values to octachlorodibenzo-p-dioxin and octachlorodibenzofuran, and removing any TEF values for all non-2,3,7,8-substituted congeners. Although it was indicated that, theoretically, it may be possible to detect nearly all of the 210 PCDD/DF isomers in the environment, seventeen 2,3,7,8-substituted congeners were known to be preferentially retained and bioaccumulated. For example, when fish or a variety of rodent species were exposed to a complex mixture of PCDDs/PCDFs from incinerator fly ash, the 2,3,7,8-substituted congeners, which were minor components of the original mixture, predominated in the analysis of their tissues (Kuehl et al., 1986; van den Berg et al., 1994). In addition, when humans were exposed to a complex mixture of more than 40 different PCDF congeners during the Oriental rice oil poisoning episodes, only the 2,3,7,8-substituted congeners were detected in subsequent blood and adipose tissue analysis (Ryan et al., 1990). EPA, which had participated in the NATO/CCMS exercise, officially adopted the revised I-TEFs in 1989, with the caveat that this risk assessment approach remains interim and continued revisions should be made (U.S. EPA, 1989; Kutz et al., 1990). The use of the TEF model for risk assessment and risk management purposes has been formally adopted by a number of countries (Canada, Germany, Italy, the Netherlands, Sweden, the United Kingdom, U.S.A.) (Yrjänheiki, 1992), and as guidance by international organizations such as the International Programme on Chemical Safety, WHO.

9.2.2. TEFs for PCBs

During the period of TEF development for PCDDs/PCDFs, a considerable body of experimental evidence was also being generated regarding the structure-activity relationships between the different polychlorinated biphenyl homologue classes (Safe, 1990, 1994). Following the synthesis of analytical standards for all 209 theoretical PCB congeners by 1984, subsequent analysis of a variety of commercial samples was able to identify all but 26 (Jones, 1988). However, once released into the environment, PCBs are subject to a variety of photolysis and biodegradation processes, to the extent that only 50-75 congeners are routinely detected in higher trophic level species (van den Berg et al., 1995). Initial structure-activity relationship studies revealed that those congeners substituted in only the meta and para positions were approximate isostereomers of TCDD. Subsequent toxicological studies confirmed that these non-ortho-substituted, "coplanar" PCBs (e.g., PCB 77, 81, 126, 169) did induce a variety of in vitro and in vivo effects similar to TCDD (Leece et al., 1985). Maximum TCDD-like activity is obtained for PCBs when there are no ortho, two or more meta, and both para positions occupied (Figure 9-1). Introduction of a single ortho substituent to the biphenyl (mono-ortho "coplanars") results in a diminishing, but not elimination, of TCDD-like activity and toxicological responses resembling commercial mixtures of PCBs. The addition of a single ortho substituent also increases the non-dioxin-like activity of the compound. Several congeners from this group are prevalent in both commercial PCBs and a wide variety of environmental samples. Some of the more persistent mono-ortho substituted PCBs (PCBs 105, 118, 156) can be found in human serum and adipose samples at levels up to three orders of magnitude higher than the "coplanar" PCBs, PCDDs and PCDFs (Patterson et al., 1994). In limited studies a third group of PCB congeners, the di-ortho "non-coplanars," has exhibited only minor amounts of dioxin-like activity (if any), usually 4-6 orders of magnitude less potent than TCDD (Safe, 1990). Recent studies have demonstrated that some of the earlier methods of preparation of these di-ortho non-coplanar PCBs had trace contaminants of PCDFs, which may account for the weak dioxin-like activity of these compounds (van der Kolk et al., 1992). In 1991, EPA convened a workshop to consider TEFs for PCBs (Barnes et al., 1991). The consensus was that a small subset of the PCBs displayed dioxin-like activity and met the criteria for inclusion in the TEF methodology. Such proposals for the TEF methodology also seem to have utility in assessing risks to wildlife (van den Berg et al., 1998; Giesey and Kannan, 1998; Ross, 2000).[1]

PCBs are often classified into two categories: "dioxin-like" and "non-dioxin-like." The dioxin-like PCBs bind to the AhR and produce dioxin-like effects in experimental animals. All other PCBs then fall into the non-dioxin-like classification. Although the dioxin-like PCBs are generally more potent at inducing biological effects, they constitute only a minor portion of the mass of PCBs found in environmental and biological samples. The non-dioxin-like PCBs account for a majority of the mass of the PCBs found in environmental and biological samples. The use of the term non-dioxin-like PCBs is not necessarily useful. The PCBs not included in the TEF scheme (i.e., the non-dioxin-like PCBs) are not a single class of compounds and have multiple toxicities with separate structure-activity relationships (Barnes et al., 1991). Not enough congener-specific research has been performed to adequately characterize or classify these compounds. For example, the "neurotoxic" PCBs have been typically defined by structure-activity relationships for decreasing dopamine concentrations or alterations in intracellular calcium in cell culture (Shain et al., 1991; Kodavanti et al., 1996).

As part of the joint WHO European Centre for Environmental Health (WHO-ECEH) and the International Programme on Chemical Safety (IPCS) project to harmonize TEF schemes for dioxin-like compounds, a database was generated consisting of all available relevant toxicological data for PCBs up to the end of 1993. Of almost 1,200 peer-reviewed publications, 146 were selected and analyzed on the basis of the following criteria: at least one PCB congener was investigated; TCDD or a reference coplanar PCB (77, 126, 169) was used during the experiment or results were available from previous experiments (same author, laboratory, experimental design); and the endpoint in question was affected by both the reference compound and the PCB congener in question (i.e., dioxin-specific). TEFs were then determined from a total of 60 articles/manuscripts on the basis of the reported results for 14 different biological/toxicological parameters. Following scientific consultation by 12 experts from 8 different countries, interim TEF values were recommended for 13 dioxin-like PCBs (Table 9-2), based on four inclusion criteria: (1) the compound should show structural similarity to PCDDs and PCDFs; (2) it should bind to the Ah receptor; (3) it should induce dioxin-specific biochemical and toxic responses; and (4) it should be persistent and accumulate in the food chain (Ahlborg et al., 1994). Increased consideration was given to selection of a TEF value based on repeat-dosing in vivo experiments, when available.

There is experimental evidence to suggest that a limited number of PCB congeners classified as weak or non-AhR agonists could effect concentration-dependent nonadditive interactions with dioxin-like compounds (Safe, 1990; 1994). Both antagonistic (Safe, 1990; Morrissey et al., 1992; Smialowicz et al., 1997b) and synergistic (Safe, 1990; van Birgelen et al., 1996a,b; van Birgelen et al., 1997) interactions between TCDD and PCBs have been observed in experimental systems. The non-additive interactions of the PCBs are thought to be mediated through non-AhR pathways (Smialowicz et al., 1997). These interactions usually occur at extremely high doses of the PCBs that are not environmentally relevant, and thus the nonadditive interactions are thought not to significantly detract from the TEF methodology (van den Berg et al., 1998; Birnbaum, 1999).

9.2.3. The Most Recent Evaluation of TEFs for PCDDs, PCDFs, and PCBs

An additional recommendation from the first WHO PCB TEF consultation was that the current database should be expanded to include all relevant information on PCDDs, PCDFs, and other dioxin-like compounds that satisfied the four inclusion criteria. Prior to the second WHOECEH consultation in 1997, various terminologies or definitions applicable to TEFs were reviewed and standardized. Whereas previously the term TEF had been used to describe all scientific endpoints used in comparison with TCDD, it was noted that a variety of experimental parameters may not be considered "toxic," but are considered as biological/biochemical responses, such as AhR binding and alkoxyresorufin O-dealkylase induction. The decision was that any experimental endpoint for which a numerical value of the relative potency compared to TCDD had been generated from a single laboratory examining a single endpoint would be known as a relative potency value, or REP. The term TEF would then be restricted to describe an order-of-magnitude consensus estimate of the toxicity of a compound relative to the toxicity of TCDD that is derived using careful scientific judgment of all available data (van Leeuwen, 1997; van den Berg et al., 1998).[1]

At the second WHO-ECEH consultation in 1997, relative potency factors were calculated based on the following methodology (van den Berg et al., 1998):

  • Assigned as reported in the publication/manuscript (verified from available data). !Calculated from the dose-response curves using linear interpolation of log doses comparing the same effect levels with correction for different control levels.
  • Calculated from ratios of low or no observed effect levels (LOELs, NOELs) and effect concentration/dose 10%, 25% or 50% values (ED/EC10,25,50).
  • Calculated from ratios of tumor promotion indexes or maximal enzyme induction levels.
  • Calculated from ratios of Ah receptor binding affinities (Kd ).

Whereas the resulting range of in vitro/in vivo REP values for a particular congener may span 3-4 orders of magnitude, final selection of a TEF value gave greater weight to REPs from repeat-dose in vivo experiments (chronic > subchronic > subacute > acute). As with the PCB TEF consultation, dioxin-specific endpoints were also given higher priority. A rounding-off procedure (nearest 1 or 5) was also employed for final TEF selection (Table 9-2). It should be noted that the TEF was rounded up or down depending on the compound, the data, and scientific judgment.

Notable amendments to the previous NATO/WHO TEF schemes include:

  • On the basis of new REPs from in vivo tumor promotion and enzyme induction, a TEF of 1.0 was recommended for 1,2,3,7,8-PeCDD.
  • Originally the TEF for OCDD was based on body burdens of the compound following subchronic exposures; a TEF based on administered dose is reduced to 0.0001.
  • New in vivo enzyme induction potency and structural similarity with OCDD support the TEF change to 0.0001 for OCDF.
  • REPs from an in vivo subchronic toxicity study (enzyme induction, hepatic retinol decreases) support reducing the TEF to 0.0001 for PCB 77.
  • A TEF value of 0.0001 was assigned for PCB 81. Even though PCB 81 was not assigned a TEF value at the 1993 WHO consultation because of lack of human residue and experimental data, more recent data demonstrate similar qualitative structural activity results compared to PCB 77.
  • Because of the lack of in vivo enzyme induction (CYP 1A1/A2) and reproductive toxicity with structurally similar congeners (PCB 47 and PCB 153), the previous interim TEF values for the di-ortho-substituted PCBs 170 and 180 were withdrawn.

Although a number of uncertainties associated with the TEF concept have been identified (nonadditive interactions with non-dioxin-like PCBs, natural ligands for the Ah receptor, questionable low-dose linearity of REP responses), the 1997 WHO expert meeting decided that an additive TEF model remained the most feasible risk assessment method for complex mixtures of dioxin-like PHAHs.

The WHO working group acknowledged that there are a number of other classes of chemicals that bind and activate the Ah receptor. The chemicals include, but are not limited to, polyhalogenated naphthalenes, diphenyl ethers, fluorenes, biphenyl methanes, quaterphenyls, and others. In addition, a number of brominated and chloro/bromo-substituted dioxin analogues of the PCDDs and PCDFs have been demonstrated to cause dioxin-like effects. The WHO working group concluded that "at present, insufficient environmental and toxicological data are available to establish a TEF value for any of the above compounds" (van den Berg et al., 1998).

The development and refinement of the TEF methodology can be thought of as an iterative process. As we accumulate more data on the biological effects of dioxin-like chemicals and a better knowledge base of their mode of action, the TEF methodology is improved. The latest evaluation of the TEF methodology for use in human health risk assessment by the WHO working group provides the most accurate assessment of the TEFs for dioxin-like chemicals. The WHO98 TEF values are recommended for use in human health risk assessment.

In January 1998, EPA and the U.S. Fish and Wildlife Service sponsored a meeting entitled "Workshop on the Application of 2,3,7,8-TCDD Toxicity Equivalency Factors to Fish and Wildlife." The major objective of the workshop was to address uncertainties associated with the use of the TEF methodology in ecological risk assessment. Twenty-one experts from academia, government, industry, and environmental groups participated in the workshop. The consensus of the workgroup was that while there are uncertainties in the TEF methodology, the use of this method decreases the overall uncertainty in the risk assessment process. However, quantifying the decrease in the uncertainty of a risk assessment using the TEF methodology remains ambiguous, as does the exact uncertainty in the TEF methodology itself (U.S. EPA, 2001).

This first section has outlined the process of assessing the relative potency of chemicals and the assignment of a consensus TEF value. There are still many questions on the use of the TEF method and the validity of some of the underlying assumptions. A detailed discussion and review of the data supporting the development and use of the TEF method, as well as the data relating to the issue of additivity, is included within the specific issues section that follows.



  1. 1.0 1.1 1.2 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.