REVIEW
The 2005 World Health Organization Reevaluation of Human and
Mammalian Toxic Equivalency Factors for Dioxins and
Dioxin-Like Compounds
Martin Van den Berg,a,1Linda S. Birnbaum,bMichael Denison,cMike De Vito,bWilliam Farland,dMark Feeley,e
Heidelore Fiedler,fHelen Hakansson,gAnnika Hanberg,gLaurie Haws,hMartin Rose,iStephen Safe,jDieter Schrenk,k
Chiharu Tohyama,lAngelika Tritscher,mJouko Tuomisto,nMats Tysklind,oNigel Walker,pand Richard E. Petersonq
aWorld Health Organization Collaborating Centre for Research on Environmental Health Risk Assessment and Institute for Risk Assessment
Sciences, Faculties of Veterinary Medicine, Science and University Medical Center, Universiteit Utrecht, PO Box 80177, 3508 TD Utrecht,
The Netherlands;bNational Health & Environmental Effects Research Laboratory, United States Environmental Protection Agency Research Triangle
Park, North Carolina 27709;cDepartment of Environmental Toxicology, University of California at Davis, Davis, California 95616-8501;dOfﬁce of
Research and Development, U.S. Environmental Protection Agency (EPA), NW, Washington, District of Columbia 20460;eChemical Health Hazard
Assessment Division, Bureau of Chemical Safety, Health Canada, Tunney’s Pasture, Ottawa, Ontario K1A OL2, Canada;fUnited Nations Environment
Program Chemicals, International Environment House, CH-1219 Cha ˆtelaine (GE), Switzerland;gInstitute of Environmental Medicine, Karolinska
Institutet, Unit of Environmental Health Risk Assessment, S-171 77 Stockholm, Sweden;hChemRisk, Austin, Texas;iCentral Science Laboratory, Sand
Hutton, YO41 1LZ York, United Kingdom;jVeterinary Physiology and Pharmacology, Texas A&M University, Texas 77843-4466;kDepartment of Food
Chemistry and Environmental Toxicology, University of Kaiserslautern, Kaiserslautern 67663, Germany;lDivision of Environmental Health Sciences,
Center for Disease Biology and Integrative Medicine, Graduate School of Medicine, The University of Tokyo, 7-3-1 Hongo, Bunkyo-ku, Tokyo 113-0033,
Japan;mInternational Programme on Chemical Safety, World Health Organization, 1211 Geneva 27, Switzerland;nNational Public Health Institute,
Department of Environmental Health, FI-70701 Kuopio, Finland;oEnvironmental Chemistry, Umea ˚University, SE-901 87 Sweden;pNational Institute
of Environmental Health Sciences, Research Triangle Park, North Carolina 27709; andqSchool of Pharmacy and Molecular and Environmental
Toxicology Center, University of Wisconsin, Madison, Wisconsin, 53705, USA
Received April 10, 2006; accepted May 20, 2006
In June 2005, a World Health Organization (WHO)-
International Programme on Chemical Safety expert meeting washeld in Geneva during which the toxic equivalency factors (TEFs) for
dioxin-like compounds, including some polychlorinated biphenyls
(PCBs), were reevaluated. For this reevaluation process, the reﬁnedTEF database recently published by Haws et al. (2006, Toxicol. Sci.
89, 4–30) was used as a starting point. Decisions about a TEF value
were made based on a combination of unweighted relative effectpotency (REP) distributions from this database, expert judgment,and point estimates. 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 of 0.03, 0.1, 0.3,etc. Changes were decided by the expert panel for 2,3,4,7,8-penta-
chlorodibenzofuran (PeCDF) (TEF ¼0.3), 1,2,3,7,8-pentachlorodi-
benzofuran (PeCDF) (TEF ¼0.03), octachlorodibenzo- p-dioxin and
octachlorodibenzofuran (TEFs ¼0.0003), 3,4,4 #,5-tetrachlorbiphen-
yl (PCB 81) (TEF ¼0.0003), 3,3 #,4,4#,5,5#-hexachlorobiphenyl (PCB169) (TEF ¼0.03), and a single TEF value (0.00003) for all relevant
mono- ortho –substituted PCBs. Additivity, an important prerequisite
of the TEF concept was again conﬁrmed by results from recent in vivo
mixture studies. Some experimental evidence shows that non-dioxin–like aryl hydrocarbon receptor agonists/antagonists are able to
impact the overall toxic potency of 2,3,7,8-tetrachlorodibenzo- p-
dioxin (TCDD) and related compounds, and this needs to be in-vestigated further. Certain individual and groups of compounds wereidentiﬁed for possible future inclusion in the TEF concept, including
3,4,4 #-TCB (PCB 37), polybrominated dibenzo- p-dioxins and diben-
zofurans, mixed polyhalogenated dibenzo- p-dioxins and dibenzofur-
ans, polyhalogenated naphthalenes, and polybrominated biphenyls.
Concern was expressed about direct application of the TEF/total
toxic equivalency (TEQ) approach to abiotic matrices, such as soil,sediment, etc., for direct application in human risk assessment. This
is problematic as the present TEF scheme and TEQ methodology are
primarily intended for estimating exposure and risks via oral inges-tion (e.g., by dietary intake). A number of future approaches todetermine alternative or additional TEFs were also identiﬁed. These
included the use of a probabilistic methodology to determine TEFs
that better describe the associated levels of uncertainty and ‘‘sys-temic’’ TEFs for blood and adipose tissue and TEQ for body burden.
Key Words: dioxins; dibenzofurans; PCBs; TEFs; reevaluation;
WHO.
Disclaimer: The contents of this paper reﬂect the opinions and views of the
authors and do not necessarily represent the ofﬁcial views or policies of
National Institute of Environmental Health Sciences, National Institutes ofHealth, USEPA, UNEP, or WHO. The mention of trade names and commercialproducts does not constitute endorsement or use recommendation.
1To whom correspondence should be addressed. E-mail: m.vandenberg@
iras.uu.nl. Fax: þ31-30-2535077.
/C211The Author 2006. Published by Oxford University Press on behalf of the Society of Toxicology. All rights reserved.
For Permissions, please email: journals.permissions@oxfordjournals.orgTOXICOLOGICAL SCIENCES 93(2) , 223–241 (2006)
doi:10.1093/toxsci/kﬂ055
Advance Access publication July 7, 2006Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
Polychlorinated dibenzo- p-dioxins (PCDDs), polychlori-
nated dibenzofurans (PCDFs), and polychlorinated biphenyls(PCBs) are persistent organic pollutants that are omnipresent inthe global environment. Many of these hydrophobic andlipophilic compounds are highly resistant to metabolism in
vertebrate species, including humans. As a result of these
properties, biomagniﬁcation occurs through the food chain, andhigh tissue concentrations can often occur in top predatorspecies. Most, if not all, toxic and biological effects of thesecompounds are mediated through the aryl hydrocarbon re-ceptor (AhR), a cytosolic receptor protein present in mostvertebrate tissues with high afﬁnity for 2,3,7,8–substitutedPCDD/Fs and some non- ortho –substituted PCBs (Poland et al. ,
1985; Safe, 1986; Safe et al. , 1985). Hundreds of congeners are
formed during synthetic processes such as combustion andcertain industrial activities (Hutzinger et al. , 1985). Thus,
human exposure either through food or the environment resultsin the uptake of a large number of these compounds. Asa result, humans retain dozens of PCB congeners in theirtissues, blood, and milk (Liem et al. , 2000; Schecter et al. ,
1994). Most PCDD and PCDF congeners with a 2,3,7,8-
chlorine substitution pattern are also strongly retained (Van den
Berg et al. , 1994). Thus, risk assessment of these compounds
involves a complex mixture of PCDD, PCDF, and PCBcompounds that are AhR agonists sharing a common mecha-nism of action and should not be done for only one speciﬁccongener.
During the last few decades, data from many experimental
studies with mixtures of these compounds are consistent with
an additive model, although deviations up to a factor of two,
and sometimes more, from additivity are not uncommon(Barnes, 1991; Barnes et al. , 1991; Birnbaum and DeVito,
1995; Safe, 1986, 1997, 1998; Safe et al. , 1985; Van den Berg
et al. , 1998; Zabel et al. , 1995). As a result of this generally
accepted additivity, the toxic equivalency concept was de-veloped during the mid 1980’s (Barnes, 1991; Barnes et al. ,
1991; Safe, 1986; Safe et al. , 1985). It uses the relative
effect potency (REP) determined for individual PCDD, PCDF,
and PCB compounds for producing toxic or biologicaleffects relative to a reference compound, usually 2,3,7,8-tetrachlorodibenzo- p-dioxin (TCDD). The total toxic equiva-
lent (TEQ) is operationally deﬁned by the sum of the productsof the concentration of each compound multiplied by its TEFvalue and is an estimate of the total 2,3,7,8-TCDD–like activityof the mixture.
Since the early 1990’s, the World Health Organization
(WHO) has organized expert meetings with the objective toharmonize the toxic equivalency factors (TEFs) for dioxin anddioxin-like compounds on the international level, therebygiving recommendations to national regulatory authorities.Prior to 2005, two WHO (re)evaluations of the TEFs wereconducted. In 1993, the ﬁrst evaluation was done that resultedin human and mammalian WHO TEFs for all 2,3,7,8-PCDDs
and PCDFs but also a recommended TEF value for severalPCBs (Ahlborg et al. , 1994). A WHO TEF (re)evaluation was
again done in 1997, which led to the revision of severalmammalian TEF values of important congeners and with-drawal of the di- ortho PCBs from the TEF concept for dioxin-
like compounds. In addition, the ﬁrst WHO TEF values for
birds and ﬁsh were proposed during this meeting (Van den Berg
et al. , 1998). To support this meeting, the Karolinska Institutet
in Stockholm (Sweden) prepared a database with results fromall studies for which REP values were known at that time,and they were used to determine the WHO 1998 TEF values.This REP database was recently used as a starting point tocompile a much more extensive database for REP values (Hawset al. , 2006). In June 2005, a third WHO expert meeting to
reevaluate current mammalian TEF values was held in Geneva,
Switzerland. Preceding this meeting, a 1-day public hearingtook place with stakeholders, interested parties, and membersof the expert panel, during which the panel members were ableto discuss various aspects of the TEF and TEQ concept with theparticipants and use this information during the actual re-evaluation process. Besides the reevaluation of the WHO 1998TEF values, the validity, criteria and correct use of the TEF/
TEQ concept, and methods for proper identiﬁcation of TEF
values and possible compounds for future inclusion were dis-cussed. This report presents the results of this meeting, includ-ing the TEF values that now are proposed as WHO 2005 TEFsfor human risk assessment of these compounds.
VALIDITY AND CRITERIA OF THE TEF CONCEPT
During the 2005 WHO reevaluation of the 1998 WHO TEF
values, both the general TEF concept and REP criteria were
extensively discussed. The criteria for inclusion of a compound
in the TEF concept at this meeting were similar to those used atthe two earlier WHO expert meetings (Ahlborg et al. , 1994;
Van den Berg et al. , 1998). These criteria are that for inclusion
in the TEF concept a compound must
/C15show a structural relationship to the PCDDs and PCDFs;
/C15bind to the AhR;
/C15elicit AhR-mediated biochemical and toxic responses;
/C15be persistent and accumulate in the food chain.
It was recognized that the vast amount of literature available in
this ﬁeld provides many examples of uncertainties associatedwith the determination of REPs. In addition, high variation cansometimes be found in REP values for the same congener andfor similar endpoints in different species (e.g., rats vs. mice).
The 2005 WHO reevaluation of the TEF values made
extensive use of the review and REP database of Haws et al.
(2006) in which a set of criteria was developed to identify,include, or exclude REPs for dioxin-like compounds. Extensiveconsultations between the compilers of this database with theWHO represented by M. van den Berg and R. E. Peterson took224
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place. However, it must be emphasized that for this 2005 TEF
reevaluation, the expert panel used all available REPs, eitherincluded or excluded in this database, and made their ownassessment (Haws et al. , 2006). Studies published since the
1997 reevaluation were also fully evaluated.
When reviewing the database of mammalian REPs for
dioxin-like compounds, it was observed that even for the mostthoroughly studied congeners like 2,3,4,7,8-pentachlorodiben-zofuran (PeCDF) and PCB 126, signiﬁcant gaps in knowledgeexist (Haws et al. , 2006). Reasons for signiﬁcant differences in
REPs for the same congener can be caused by the use ofdifferent dosing regimens (acute vs. subchronic), differentendpoints, species, and mechanisms (e.g., tumor promotion
caused by at least two different mechanisms as for mono-
ortho –substituted PCBs), as well as different methods used for
calculating REPs. Thus, different methodological approachesused in different studies clearly provide uncertainties whenderiving and comparing REPs. If future study designs to deriveREPs were more standardized and similar, the variation inREPs when using the same congener, endpoint, and speciesmight be expected to be smaller.
At this 2005 meeting, the ‘‘ideal’’ REP study design was
discussed as previous WHO TEF evaluations did not providesufﬁcient information regarding the criteria that needed to bemet to establish an REP value and give an expert panel thegreatest degree of conﬁdence in a particular REP. Thefollowing general guidelines for a future ideal dose-responsestudy used to determine an in vivo REP resulted from the
workshop:
/C15A full dose-response curve for both the congener and for
2,3,7,8-TCDD should be determined.
/C15The congener and 2,3,7,8-TCDD should be administered
by the same route to animals of the same species, strain, sex,
and age, and the animals should be housed, fed the same diet,and maintained under the same conditions in the samelaboratory.
/C15Ideally, the absolute maximal response (efﬁcacy) should
be similar for both the congener and for 2,3,7,8-TCDD andtheir dose-response curves should be parallel, but in practice,this is often not observed for various reasons.
/C15If the above dose-response criteria are met, the REP
should be calculated by dividing the effective dose 50%(ED
50) of 2,3,7,8-TCDD by the ED 50of the congener.
/C15If full dose-response relationships are not attained and
determination of ED 50’s is not possible, lowest observed effect
doses or concentrations or benchmark doses could be used todetermine the REP. However, such an REP has more un-certainty than if ED
50’s were used.
For studies that are designed to determine REPs, it is clear that
in vivo studies have the highest priority because they combine
both toxicokinetic as well as toxicodynamic aspects. Therefore,in vivo studies should preferably be used for setting TEFs.
Nevertheless, in vitro studies can contribute signiﬁcantly toestablish the AhR-mediated mechanism of action of a com-
pound and explain possible differences in species sensitivity,especially with respect to that of humans versus experimentalanimal species. For in vitro studies, stricter criteria should be
applied as these are from an experimental design point of view
usually easier to accommodate than in vivo studies. For in vitro
studies, the following experimental design is suggested todetermine an REP:
/C15A vehicle group and at least four graded concentrations of
a congener and four graded concentrations of 2,3,7,8-TCDDshould be selected.
/C15For congener and 2,3,7,8-TCDD treatment groups, three
of these concentrations should elicit a response that fallsbetween the EC
20and EC 80for the congener and for 2,3,7,8-
TCDD.
/C15At least one concentration should elicit a maximal
response (EC 100), and the concentration-response curves
should be parallel.
/C15The REP should be based on the EC 50of 2,3,7,8-TCDD
and the EC 50of the congener.
In general, 2,3,7,8-TCDD has been used as the reference
compound of choice, but in several studies, PCB 126 has beenused instead of 2,3,7,8-TCDD. Based on available data fromthe literature, it was concluded that PCB 126 could indeed beused as a reference compound in rat studies with an REP valueof 0.1. Recent studies have conﬁrmed this value for multipleendpoints (Toyoshiba et al. , 2004; Walker et al. , 2005).
However, it should be examined in more detail if the same
REP for PCB 126 is applicable as a reference compound for
mouse studies. The REP values for some endpoints such asenzyme induction tend to be signiﬁcantly lower in mice than inrats (Birnbaum and DeVito, 1995; DeVito et al. , 2000; Harper
et al. , 1993; van Birgelen et al. , 1996a). In this respect, it
should be noted that mice studies in which PCB 126 was usedas a reference compound were excluded from the database andfrom further consideration because of other methodological
reasons (Haws et al. , 2006).
Literature data also indicate that the PCB 126 REP for
enzyme induction in human cell systems, including primaryhepatocytes, breast cancer cell lines, and primary lymphocytes,may be one or two orders of magnitude lower (van Duursenet al. , 2003; Zeiger et al. , 2001). In addition, the apparent
binding afﬁnity of 2,3,7,8-TCDD to the human AhR isgenerally 1/10th that of the AhR of the more sensitive rodent
species, but signiﬁcant variation among individual humans
occurs (Ema et al. , 1994; Harper et al. , 2002; Poland et al. ,
1994; Ramadoss and Perdew, 2004; Roberts et al. , 1990). It has
been suggested that on average, humans are among the moredioxin-resistant species, but the human data set is too limited tobe conclusive (Harper et al. , 2002; Okey et al. , 2005). A study
with AhR-humanized mice may indicate lower responsivenesstoward toxic effects of 2,3,7,8-TCDD (Moriguchi et al. , 2003).
Taken together, this information warrants more research into2005 WHO REEV ALUATION OF TEFs 225Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
REP values in human systems to establish if the present TEFs
based on rodent studies are indeed also valid for humans.
Additivity is an important prerequisite of the TEF concept,
and this aspect was revisited in detail by the 2005 expert panel.It was concluded that results from recent in vivo mixture
studies with dioxin-like compounds are consistent with addi-
tivity and support the TEF approach (Fattore et al. , 2000; Gao
et al. , 1999; Hamm et al. , 2003; Walker et al. , 2005). Gao et al.
(1999) studied the REP and additivity of 2,3,7,8-TCDD,1,2,3,7,8-PeCDD, and 1,2,3,4,7,8-HxCDD in a rat ovulationmodel; their results conﬁrmed both parallel dose-responsecurves and mixture additivity. Fattore et al. (2000) measured
hepatic vitamin A reduction in rats after subchronic dietary
exposure to a low-dose mixture containing 1,2,3,7,8-PeCDD,
2,3,4,7,8-PeCDF, and 1,2,3,6,7,8-HxCDF to test additivity. Theeffects of this mixture showed that the predicted results basedon WHO 1998 TEFs were approximately twofold higher.Hamm et al. (2003) studied a mixture of nine dioxins, furans,
and coplanar PCBs and looked at developmental reproductiveendpoints in rats, comparing results of the mixture to that of2,3,7,8-TCDD alone. The results showed that the experimental
estimated TEQ was within a factor of two of that predicted
from the WHO 1998 TEFs. A mixture study from the NationalToxicology Program was also examined by the expert panel,and again the results generally supported additivity and paralleldose-response curves for complex and long-term neoplasticand nonneoplastic endpoints (Walker et al. , 2005).
Thus, results in these recent mixture studies could be
predicted rather well with the WHO 1998 TEFs, within a factor
of two or less. This degree of accuracy was somewhat
surprising in view of the complicated experimental situationpresent in subchronic toxicity studies, where congener-speciﬁctoxicodynamics and kinetics are intermingled and can in-ﬂuence the ﬁnal outcome. In addition, the WHO 1998 TEFswere derived from a range of REPs using different biologicalmodels or endpoints and were therefore estimates with an orderof magnitude uncertainty (Van den Berg et al. , 1998).
PROCESS USED TO DETERMINE TEF VALUES:
POINT ESTIMATES, EXPERT JUDGMENT, AND
PROBABILISTIC DISTRIBUTION
Both the WHO 1993 and 1997 TEF reevaluations used point
estimates derived by expert judgment from a wide range of
REPs (Ahlborg et al. , 1994; Van den Berg et al. , 1998). In the
2005 TEF reevaluation, it was decided by the expert panel touse the REP database from Haws et al . (2006) for initial
assessment of a TEF value. This recently published databaseand applied criteria were a reﬁnement of the criteria anddatabase that were developed to support the two previous WHOTEF reevaluations (Ahlborg et al. , 1994; Van den Berg et al. ,
1998). The criteria for inclusion or exclusion of an REP in thisdatabase (Haws et al. , 2006) were accepted by the expert panel.
These criteria can be summarized as follows:
/C15At least one test congener and a valid reference
compound must have been included in the study or thereference compound must have been included in an identicalexperiment from the same laboratory, but in another study.
/C15The endpoint must have been an established AhR-
mediated response known to be affected by both the test
congener and the reference compound.
/C15In the REP database, in vivo and in vitro studies were
separated.
/C15Repetitive endpoints (i.e., measures of the same bi-
ological response) were identiﬁed in all studies in thedatabase, and the most representative REP value was retainedfor reevaluation of a TEF.
/C15Those studies that used only a single-dose level of either
the test and/or reference compound were ﬁltered out of the
REP database and not used in the TEF reevaluation process.
/C15Results from non-peer–reviewed studies were not used in
reevaluating a TEF value and consequently did not contributeto the distribution of REPs for individual congeners.
/C15REPs based on biological responses that were statistically
signiﬁcant were included in the 2005 REP database andcontributed to the distribution of REPs for individual con-
geners used to reevaluate TEFs. However, when there was
a very limited data set for an individual congener, the panelalso considered biological responses that were not statisticallysigniﬁcant as part of the overall expert judgment in reevaluat-ing a TEF value.
/C15REPs based on quantitative structure-activity relationship
studies were included in the REP database.
When using this database, the primary focus of the TEF
reevaluation was on in vivo studies (Haws et al. , 2006). In vitro
studies were only used for support in those situations where noor very few in vivo REP data were available. For in vitro REPs,
only established AhR-mediated responses were used to assign
REP values.
During the TEF reevaluation, the expert panel considered
using REP distributions available from the REP database(Haws et al. , 2006) when reevaluating a TEF value. However,
the REP distributions in this study are unweighted (Haws et al. ,
2006), and it was decided that establishing a weighting criteriafor REPs generated in different types of studies ( in vivo ,
in vitro , chronic, acute, etc.) was not feasible at this meeting. In
addition, it was concluded that REP distributions for a speciﬁc
congener in this database could not be used to derive a TEFvalue because a ﬁxed percentile would have to be used asa cutoff point. Such an approach would be like using a singlepoint estimate, but with lower biological or toxicologicalrelevance. This is because all types of in vivo studies (acute,
subchronic, etc.) and different endpoints have been combined,and associated REP distributions are shown as a single box plot.
Thus, with only unweighted distributions of REPs available,226
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a ﬁnal expert judgment in the TEF reevaluation process
involving the type and quality of the study had preferenceover the unweighted REP distributions (Haws et al. , 2006).
Nevertheless, it was recognized by the expert panel that inthe future weighted REP distributions could be used for
derivation of TEF values, but establishing consensus values
for these REP weighting factors would require additionalexpertise.
The WHO expert panel decided that a combination of these
unweighted REP distributions, expert judgment, and pointestimates would be used to reevaluate a TEF. Figure 1 showsthe unweighted distribution of in vivo andin vitro REPs and
WHO 1998 TEF values for PCDDs, PCDFs, non- ortho PCBs,
and mono- ortho PCBs (Haws et al. , 2006). These unweighted
REP distributions were used to start the selection and decisionprocess for a TEF reevaluation. The 75th percentile of thein vivo REP distribution for an individual congener was used as
an initial decision point to review the WHO 1998 TEF for thatcongener. If the WHO 1998 TEF was below the 75th percentileof the in vivo REP distribution, the data driving this TEF value
was extensively reevaluated. If the WHO 1998 TEF value was
above the 75th percentile, a quick review was done regarding
the decision made at the 1997 WHO meeting with respect tothose studies that had been driving the 1998 WHO TEF value.In addition, results of new studies conducted after 1997 or oldinformation missed at the 1997 WHO meeting were evaluatedto determine if these would inﬂuence the WHO 1998 TEFvalue for that congener. Based on the combined information,a possible new TEF value was considered. Special attention
was also given to validity of WHO 1998 TEF values that were
near or higher than the 90th percentile, e.g., 1,2,3,7,8-PeCDF.Thus, the above TEF reevaluation process provided a way bothto increase as well as to decrease a TEF value. Figure 2illustrates the decision scheme used at the expert meeting forthe initial reevaluation process of the TEFs. For transparency,the expert judgment process and rationale used by the expertpanel for a possible newly assigned WHO 2005 TEF value is
explained in the next paragraph. This is followed by subsequent
paragraphs devoted exclusively to each congener reevaluated.
As in previous WHO TEF consultations, it was decided by
the expert panel to use a stepwise scale for assigning TEFsvalues. However, instead of assigning TEFs in the incrementsused previously (0.01, 0.05, 0.1, etc.), it was decided to use halforder of magnitude increments on a logarithmic scale at 0.03,0.1, 0.3, etc. As a result, all (non)revised 2005 WHO TEFs
were ﬁtted on a logarithmic scale. This decision to assign TEFs
as half order of magnitude estimates may be useful indescribing, with statistical methods, the uncertainty of TEFsin the future. Thus, as a default, all TEF values are assumed tovary in uncertainty by at least one order of magnitude,depending on the congener and its REP distribution. Conse-quently, a TEF of 0.1 infers a degree of uncertainty bounded by0.03 and 0.3. For a TEF value of 0.3, a degree of uncertainty
bounded by 0.1 and 1 was used. Thus, the TEF is a central valuewith a degree of uncertainty assumed to be at least ± half a log,
which is one order of magnitude. However, it should berealized that TEF assignments are usually within the 50th to75th percentile of the REP distribution, with a general in-clination toward the 75th percentile in order to be health
protective. However, the latter approach was also inﬂuenced by
the type and quality of study, e.g., single versus multiple dose,that could not been discerned from the REP distributions shownin Figure 1. This more conservative and health protectiveapproach practically means that for a TEF value the likelihoodof a half-log error too low is less than the likelihood of halfa log error too high. Due to the new ‘‘spacing’’ to express TEFson a half-log scale, it was also necessary in the ﬁnal review
process to evaluate each individual TEF value for those
congeners for which there were no new data available.
WHO 2005 TEF VALUES
1,2,3,7,8-PeCDD
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 provideREPs 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.
1,2,3,4,7,8-HxCDD
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 (Gaoet 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 bechanged to either 0.3 or 0.03. Therefore, it was decided to keepthe WHO 2005 TEF value at 0.1.
1,2,3,6,7,8-HxCDD
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 WHO2005 TEF was decided.
1,2,3,7,8,9-HxCDD
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 are2005 WHO REEV ALUATION OF TEFs 227Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
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 panelconsidered decreasing the TEF value to 0.03 but decided thatthere 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 wastherefore decided to retain the old value of 0.1 as the WHO2005 TEF value.
FIG. 1. Distribution of in vivo unweighted REP values in the REP 2004database. Reprinted with permission from (Haws et al. , 2006).228 VAN DEN BERG ET AL.Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
1,2,3,4,6,7,8-HpCDD
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 thiscongener (Simanainen et al. , 2002; Viluksela et al. , 1994,
1997a,b). An earlier tumor promotion study also indicateda 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 theTEF to 0.03, which is well above the 90th percentile of the REPdistribution. This suggestion was rejected by the expert panel.It was decided to retain the WHO 2005 TEF value as 0.01.
Octachlorodibenzo- p-dioxin
The WHO 1998 TEF was set at 0.0001, which is well outside
the 10th percentile of the range of in vivo andin 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 theresults 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 morerecent 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 con-
taminated with other more potent 2,3,7,8–substituted conge-ners such as 2,3,4,7,8-PeCDF. Using the NTP data nowavailable 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 theresults (Calculation of 2,3,4,7,8-PeCDF contamination inOCDD Fattore et al. study [2000]. 2.5 ppm ¼2.5 pg/ lg
OCDD. Highest OCDD dose 800 ppb ¼800 ng/g ¼0.8lg/g
feed. At this dose level, the PeCDF dose must have been 2.5 pgPeCDF/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, whichis equivalent with 200 pg PeCDF/kg/day or 0.2 ng PeCDF/kg/day. This dose is two orders of magnitude lower than the lowestdose (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 concludedthat there is very limited in vivo information available with only
one subchronic toxicity study (Fattore et al. , 2000; Werme-
linger 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 increasethe TEF value for this compound to 0.001, but the combined
information from in vivo andin vitro data (Haws et al. , 2006)
did justify a raise in TEF value. Therefore, it was decided to2005 Database
98 WHO TEFsBasis:
> 75 percentile of
REP distribution
Yes No
Quick review of 97
decisionCheck in database
what drives TEF 
Evaluate new data
-Define confidence
-justification *Evaluate new data if 
impact TEF
Yes No
-Define confidence
-justification *
* indicate where TEF falls within REP
distribution, and give clear rational for 
decision Derive TEF
-Define confidence
-justification * Derive TEF Confirm current TEF
FIG. 2. Decision scheme used in the 2005 reevaluation of the 1998 WHO TEF values (Van den Berg et al. , 1998) assigned to individual PCDD, PCDF, and
PCB congeners.2005 WHO REEV ALUATION OF TEFs 229Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
increase the WHO 2005 TEF value to 0.0003. The expert panel
is aware of the implications that the increase in this WHO TEFvalue for OCDD might have from a regulatory and riskmanagement point of view. However, with respect to the highconcentrations 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 ofTEQ for abiotic environmental matrices.
2,3,7,8-TCDF
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 TEFshould remain at 0.1.
1,2,3,7,8-PeCDF
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 shouldbecome 0.03.
2,3,4,7,8-PeCDF
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 availableto evaluate this earlier TEF value more closely. The REPs forneoplastic endpoints from the NTP study (Walker et al. , 2005)
are around 0.2–0.3, while nonneoplastic endpoints have REPsthat range from 0.7 to 1.1. An earlier subchronic study byPluess 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 panelto change the WHO 2005 TEF to 0.3.
1,2,3,4,7,8-HxCDF
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 andin vivo studies with
the PCDF are combined, REP values lie within the 50th and75th percentile. A new study reported an REP of 0.03 forhepatic vitamin A reduction (Fattore et al. , 2000). However, the
animals analyzed were from an earlier study from which anREP 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 allaround the 50th to 75th percentile (Haws et al. , 2006), the expert
panel decided not to discriminate between TEF values for thesecongeners. As a result, the 2005 WHO TEF remains at 0.1.
1,2,3,7,8,9-HxCDF
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.
2,3,4,6,7,8-HxCDF
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 inresponses for this and certain other HxCDFs, there was con-sensus 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 for1,2,3,4,7,8,9-HpCDF. Although these in vitro results do suggest
a slightly higher TEF than 0.01, the expert panel thought thatthere was too much uncertainty in this limited database to raisethe 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.
Octachlorodibenzofuran
The WHO 1998 TEF value of 0.0001 is within the 50th and
75th percentile of the REP distribution range of three in vivo230
VAN DEN BERG ET AL.Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
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 octachlorodi-
benzofuran (OCDF) greater than 0.0001 based on hepatic
vitamin A reductions. Some earlier in vivo studies also indi-
cated 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 panelabout 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 andin vitro REPs ( <10) are reviewed, REPs
range from 4 310
/C2556to 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, butincreasing the WHO 2005 TEF value to 0.0003 is appropriatein 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).
PCB 77
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 aroundthe 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 mostrepresentative (Chu et al. , 1995). As a consequence, the WHO
2005 TEF value remained at 0.0001.
PCB 81
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 TEFscheme. 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 monkeyhepatocytes indicate that PCB 81 is more potent than PCB77 (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 lowend 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 toraise the WHO 2005 TEF value to 0.0003. However, the expert
panel expressed its low conﬁdence in the PCB 81 REP databasebecause it lacks in vivo REP data.
PCB 126
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 withthis 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 thiscongener (Haws et al. , 2006), which supports the use of PCB
126 as reference compound with a TEF of 0.1 when comparingrat 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 thisissue. Although concern was expressed about interspeciesvariability in REPs, the expert panel considered the presentinformation too limited to make a decision other than to retain0.1 as the WHO 2005 TEF.
PCB 169
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 dosemouse 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 andendpoints. In view of the fact that the WHO 1998 TEF wasalso 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 tohave signiﬁcantly higher REPs (around 0.1) than thoseobserved in a 13-week study. In view of these signiﬁcantdifferences between single- and multiple-dose studies, theexpert panel judged that the WHO 2005 TEF for PCB 169 of0.03 would be more appropriate than a potentially overlyconservative 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 four2005 WHO REEV ALUATION OF TEFs 231Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
to ﬁve orders of magnitude, depending on the congener. In
Figure 3, this wide variation in REPs is illustrated. Even if onlyin vivo studies are considered, the 90% distribution range is
extremely large (see Figure 1). This great variation in REPvalues 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 com-
pounds to play, at least in part, a role in causing this largevariation. De Vito (2003) found that less than 1% contamina-tion of PCB 77 by PCB 126 signiﬁcantly impacted the apparentREP of PCB 77. Shortly before the WHO 2005 TEF reevalua-tion meeting, two laboratories of panel members performeda 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 CYP1A1induction in two genetically modiﬁed rodent hepatoma CA-FLUX cell lines at concentrations that would generally justifyan REP larger than 0.0001. Based on the combined informa-tion, 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 mono-ortho PCBs can probably be explained by the occurrence of
impurities with 2,3,7,8-chlorine–substituted PCDDs andPCDFs or PCB 126. As the occurrence of these impuritiesclearly depends on the route of synthesis and the degree ofcleanup, it was not possible to make a general statement abouthow 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 establisha 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 decisionscheme with /C2175th 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 threemono- 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. Conse-quently, 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
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 foodchain (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, noinduction of CYP1A1 or 1B1 could be found. However, PCB37 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 andin 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 poly-
brominated dibenzo- p-dioxins (PBDDs) and polybrominated
dibenzofurans (PBDFs) have AhR agonist properties and causedioxin-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 showingthat these compounds are found in human milk and adiposetissue at levels that can contribute signiﬁcantly to the totalamount of TEQ (Choi et al. , 2003; Kotz et al. , 2005; Ohta et al. ,
2005). It appears that environmental levels might be signiﬁ-
cantly lower than those of the PCDDs, PCDFs, and PCBs
already in the TEF scheme. However, a better exposureassessment especially with regard to humans is needed. If thepresence of PBDDs and PBDFs in human food as well as inpeople 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.232 VAN DEN BERG ET AL.Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
need for assigning TEFs to these compounds. At present, it is
unclear to what extent the ongoing use of brominated ﬂameretardants, especially polybrominated diphenylethers (PBDEs),could lead to an increase in human and environmental exposureto 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 com-pounds proposed in this paragraph for development of WHOTEFs, 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) dioxinsand dibenzofurans are indeed detected in humans and theirfood, these should deﬁnitely be considered for inclusion in theTEF scheme. Early in vitro studies suggest that these com-
pounds follow the same structure-activity rules as the PCDDsand PCDFs (Behnisch et al. , 2001; Mason et al. , 1987; Weber
and Greim, 1997).
Hexachlorobenzene
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 beconﬁrmed that highly puriﬁed HCB has indeed AhR agonisticproperties, as contamination of HCB with PCDDs and PCDFshas been reported (Goldstein, 1979) (Analysis of HCB done forthe U.K. Medical Research Council indicated levels of 16,000ng 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 studiescould have an impurity problem similar to that observed forthe mono- ortho PCBs. Priority should thus be given to conﬁrm
the compound’s dioxin-like properties using highly puriﬁedHCB with measured absence of 2,3,7,8-chlorine–substituteddioxins 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 beconsidered 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 inducedioxin-like effects, such as cleft palate and hydronephrosis
(Birnbaum et al. , 1983; McKinney and McConnell, 1982;
Miller and Birnbaum, 1986). Further arguments for inclusionwould be that multiple PCN and PBN congeners have distinctin 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.
Polybrominated Biphenyls
Certain polybrominated biphenyls (PBBs) have been re-
ported to be AhR active in both in vitro andin 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, PBBexposure has been reported in bird species at the top of the foodchain 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 furtheridentify 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 relevanceof PBBs to the total TEQ.
Polybrominated Diphenylethers
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 signiﬁ-cant 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 becauseof 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 alsoproduce PBDDs and PBDFs. In conclusion, it was recom-mended 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 (Denison2005 WHO REEV ALUATION OF TEFs 233Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
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 herbalsupplements 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 orinhibit induction of CYP1A1-dependent activities by 2,3,7,8-TCDD in cell culture model systems have been reported bynumerous 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 naturallyoccurring 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 forindole-3-carbinol (I3C) (Wilker et al. , 1996). In addition,
naturally occurring AhR ligands, such as I3C and diindoly-methane, have been reported to inhibit 2,3,7,8-TCDD–de-pendent 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 immuno-
toxicity in C57BL/6J mice has also been reported (Bannisterand 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 ondioxin toxicity of non-dioxin–like compounds, e.g., thyroidhormones, porphyrins, reproductive toxicity, and immunotox-
icity (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 theoverall toxic potency of 2,3,7,8-TCDD and related compounds.If occurring under natural background situations, these inter-actions 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 ofindividual REP or TEF values for dioxin-like chemicals. Thepotential impact of these non-dioxin–like natural compoundson the risk of toxicity posed by exposure to a particular level ofTEQs should be further investigated.
THE USE OF TEQ FOR ABIOTIC ENVIRONMENTAL
MATRICES
Concurrent with the development of the TEF and TEQ
approach has been its application to environmental matrices
such as soil, sediment, industrial wastes, soot, ﬂy ash frommunicipal incinerators, waste water efﬂuents, etc. As such, theTEQ approach has been and continues to be used to givea single value to complex environmental matrices (Barnes,1991; Barnes et al. , 1991), usually without taking into
consideration whether this is actually a risk-based number.The expert panel emphasized that correct application of thepresent TEF scheme (see Table 1) and TEQ methodology inhuman risk assessment is only intended for estimating exposure
to dioxin-like chemicals from consumption of food products,
breast milk, etc. This limitation is derived from the fact thatthose REP studies that have been considered most relevant forthe determination of the present TEFs are largely based on oralintake studies, often through the diet. In fact, experimentaltoxicological studies using abiotic matrices with dioxin-likecompounds that would allow for the determination of environ-mental matrice-based REPs (e.g., soil or sediment) are almost
nonexistent. Furthermore, the issue of matrix-speciﬁc bio-
availability of these chemicals from abiotic environmentalsamples leads to a high degree of uncertainty for riskassessment as this is largely dependent upon the organiccarbon content and age of the particles. For example, direct
TABLE 1
Summary of WHO 1998 and WHO 2005 TEF Values
Compound WHO 1998 TEF WHO 2005 TEF
Chlorinated dibenzo- p-dioxins
2,3,7,8-TCDD 1 11,2,3,7,8-PeCDD 1 1
1,2,3,4,7,8-HxCDD 0.1 0.1
1,2,3,6,7,8-HxCDD 0.1 0.11,2,3,7,8,9-HxCDD 0.1 0.11,2,3,4,6,7,8-HpCDD 0.01 0.01
OCDD 0.0001 0.0003
Chlorinated dibenzofurans
2,3,7,8-TCDF 0.1 0.11,2,3,7,8-PeCDF 0.05 0.03
2,3,4,7,8-PeCDF 0.5 0.3
1,2,3,4,7,8-HxCDF 0.1 0.11,2,3,6,7,8-HxCDF 0.1 0.1
1,2,3,7,8,9-HxCDF 0.1 0.1
2,3,4,6,7,8-HxCDF 0.1 0.11,2,3,4,6,7,8-HpCDF 0.01 0.011,2,3,4,7,8,9-HpCDF 0.01 0.01
OCDF 0.0001 0.0003
Non- ortho –substituted PCBs
3,3#,4,4#-tetraCB (PCB 77) 0.0001 0.0001
3,4,4 #,5-tetraCB (PCB 81) 0.0001 0.0003
3,3#,4,4#,5-pentaCB (PCB 126) 0.1 0.1
3,3#,4,4#,5,5#-hexaCB (PCB 169) 0.01 0.03
Mono- ortho –substituted PCBs
2,3,3 #,4,4#-pentaCB (PCB 105) 0.0001 0.00003
2,3,4,4 #,5-pentaCB (PCB 114) 0.0005 0.00003
2,3#,4,4#,5-pentaCB (PCB 118) 0.0001 0.00003
2#,3,4,4 #,5-pentaCB (PCB 123) 0.0001 0.00003
2,3,3 #,4,4#,5-hexaCB (PCB 156) 0.0005 0.00003
2,3,3 #,4,4#,5#-hexaCB (PCB 157) 0.0005 0.00003
2,3#,4,4#,5,5#-hexaCB (PCB 167) 0.00001 0.00003
2,3,3 #,4,4#,5,5#-heptaCB (PCB 189) 0.0001 0.00003
Bold values indicate a change in TEF value.234 VAN DEN BERG ET AL.Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
application of these WHO TEFs for assessment of OCDD or
OCDF present in soil, sediment, or ﬂy ash would lead toinaccurate assessment of the potential toxic potency of thematrix. This derives primarily from the fact that the highlyhydrophobic PCDDs and PCDFs bind strongly to particles
thereby signiﬁcantly reducing their bioavailability for living
organisms (Van den Berg et al. , 1994). As a result, application
of these WHO TEFs for calculating the TEQ, e.g., OCDD andOCDF, in abiotic environmental matrices has limited toxico-logical relevance and use for risk assessment unless the aspectof reduced bioavailability is taken into consideration. Never-theless, the expert panel recognized that it is now commonpractice to use the TEQ and associated TEFs directly to
characterize and compare contamination by dioxin-like chem-
icals of abiotic environmental samples and is even codiﬁed innational and international legislation, e.g., the StockholmConvention on Persistent Organic Pollutants.
In relation to this use of the TEQ, it should be emphasized
that while these values by themselves do not have anytoxicological implications or direct use in risk assessment,they can be a useful tool to compare concentrations within
similar abiotic matrices and serve a prioritization function.
Accordingly, it is recommended that when a human riskassessment is to be done from abiotic matrices, factors suchas fate, transport, and bioavailibility from each matrix bespeciﬁcally considered before a ﬁnal estimate of the toxico-logical relevant TEQ is made. If a human risk assessment isdone for abiotic matrices, the expert panel recognized that itwould be preferable to use congener-speciﬁc equations
throughout the whole model rather than base it on total TEQ
in an abiotic matrix.
FUTURE RECOMMENDATIONS FOR DETERMINATION
OF TEF VALUES
Previous WHO TEF reevaluations have used expert judg-
ment and point estimates to establish congener-speciﬁc TEFvalues. In addition, the 1997 expert meeting indicated that TEFvalues were order of magnitude estimates (Van den Berg et al. ,
1998). This statement was given irrespective of the type of
congener, even though large differences are present in the REP
studies of individual compounds (Haws et al. , 2006). When
using point estimates and expert judgment, an advantage is thatselection of a TEF can be made from those studies which aremost relevant for human exposure (e.g., in vivo long term or
subchronic). A disadvantage is that such an approach does notdescribe the range of REPs and may reﬂect a bias in judgmentwithin the expert panel.
Recently, several authors have published papers in which
they advocated the use of a probabilistic approach to determineTEFs (Finley et al. , 2003; Haws et al. , 2006). In using such an
approach, there is a clear advantage because it will betterdescribe the level of uncertainty present in a TEF value. Thedistribution of REPs can be expressed in terms of minimum and
maximum values combined with percentiles at different levels(e.g., 25th and 75th percentiles). A disadvantage could be thatsuch an approach lumps all data together and gives similarweight to all types of studies. In part, this problem could be
avoided by separating in vitro from in vivo REPs (Haws et al. ,
2006).
However, if probabilistic approaches for setting a future TEF
are used, it is essential that weighting factors be applied toREPs that are determined from different types of studies. Theseweighted REP values could then be used to determine weightedREP distributions in the risk assessment process. Clearly,unweighted REP distribution ranges that bracket the TEFs
incorporate biological and toxicological uncertainty. For this
reason, in the WHO 2005 TEF reevaluation, unweighted REPdistribution ranges, expert judgment, and point estimates wereused in combination to assign TEFs. The sole use ofa probabilistic approach to determine TEF values also includesother decision points, such as establishing a range instead ofa point estimate for the TEF value. However, the use of a TEFrange might cause problems for regulatory authorities and
international harmonization of TEF values because one or
more TEF values could then be selected for risk assessmentcalculations. This might easily lead to different TEFs beingused by different countries depending on the level of conser-vatism used in the risk management process by nationalauthorities. In this respect, the choice, e.g., of a 50th, 75th, or95th percentile of the REP distribution range to assign a TEF isa risk management decision.
Similar to the use of WHO 2005 TEFs and TEQ with abiotic
matrices, the application of these values to human tissuesamples must be carried out with caution. This is because thepresent WHO TEF concept is, by default, primarily designedfor intake situations. There is emerging evidence suggestingthat the REP of certain dioxin-like compounds may differ whenthe REP is determined based on administered dose versus tissueconcentration (Chen et al. , 2001; DeVito et al. , 2000; Hamm
et al. , 2003). As a result, the use of systemic TEFs and TEQ has
been suggested as an additional approach to the present WHOTEFs. From a biological and toxicological point of view, thedevelopment and use of systemic TEFs is recommended, butthe expert panel was of the opinion that at present there isinsufﬁcient data to allow the development of systemic TEFs. Ifsystemic TEFs would be developed in the future, TEF valuesbased on blood lipid concentration might be the preferred
choice. However, the use of intake TEFs from food is a valid
approach for estimation of human body burdens since many ofthe concerns with issues of fate and transport when dealingwith abiotic matrices do not exist and many of the pharmaco-kinetic issues are already (partially) dealt with during bio-accumulation and biomagniﬁcation up the food chain.
With respect to the use of systemic TEFs, it would also be
useful to determine if in vitro –derived TEFs can potentially be
used as surrogates for systemic TEFs derived from in vivo2005 WHO REEV ALUATION OF TEFs 235Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
studies. If such a relationship does exist, this would allow
a better use of the vast amount of in vitro data that has been
obtained for dioxin-like compounds over the last few decades.In view of their direct biological relevance to humans, theexpert panel proposed that systemic or body burden TEFs for
humans should be developed in the near future. These body
burden/systemic TEFs would allow a more accurate quantita-tive human dose-response assessment. However, it was alsoconcluded by the expert panel that such systemic TEFs shouldbe used in the future along with the 2005 WHO TEFs derivedfor ingestion situations as both types of TEFs have differentvalid applications. The TEQ based on intake TEFs can be usedto monitor intervention programs, while systemic or body
burden TEFs would be more applicable for biomonitoring
systemic levels of dioxin-like chemicals in humans. Inaddition, body burden TEFs can also be used as the dosemetric for interspecies extrapolation. At present, the WHO2005 TEFs that are based on intake can be applied forcharacterization of exposure to dioxin-like chemicals in humanblood or tissues and comparisons across populations, but thesederived TEQ values have certain caveats from a risk assess-
ment point of view.
CONCLUSIONS
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; Hamm et al., 2003;
Walker et al. , 2005). These studies showed that WHO 1998 TEF
values predicted mixture toxicity within a factor two or less.Such accuracy is almost surprising in view of the fact that TEFsare derived from a range of REPs using different biological
models or endpoints and are considered estimates with an orderof 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 onthe overall accuracy of the estimated magnitude of the TEQneeds to be investigated further, but it does not impact theexperimental determination of individual REPs or TEFs.
For this TEF reevaluation process, the expert panel made
extensive use of the reﬁned TEF database that was recentlypublished 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 solelyunweighted REP distributions to set a TEF value was rejectedbecause 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 ortoxicological 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, withstatistical methods, the uncertainty of TEFs in the future. InTable 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 bioticsamples. 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 thetotal TEQ of these samples with an overall decrease in TEQ
FIG. 4. Percent reduction in total TEQ levels calculated for the same biotic samples when 2005 TEFs rather than 1998 TEFs are used. For each biotic sample
shown, the height of the bar is the percent that the total TEQ level determined using WHO 2005 TEFs is of the total TEQ level determined using WHO 1998 TEFs .236 VAN DEN BERG ET AL.Downloaded from https://academic.oup.com/toxsci/article-abstract/93/2/223/1707690 by Subsecretaria de Salud Publica user on 05 August 2020
ranging between 10 and 25%. The exception being the chicken
where the decrease of the TEF for 2,3,4,7,8-PeCDF (from 0.5to 0.3) and of lower TEFs for the mono- ortho PCBs resulted in
an almost 50% decrease of total TEQ. In view of this averageimpact 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 mechanis-tic considerations, PCB 37, PBDDs, PBDFs, PXCDDs,PXCDFs, PCNs, PBNs, and PBBs undoubtedly belong in theTEF concept. However, for most, if not all, of these compoundsthere 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 humanswith respect to TEQ dietary intake. In addition, HCB could bea possible candidate for inclusion in the TEF/TEQ concept butonly if it is unequivocally shown that impurities have not beenthe cause of earlier dioxin-like effects observed in experimentalmodels. With respect to PBDEs, it was concluded that there isno 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) andTEQ methodology are primarily meant for estimating exposurevia dietary intake situations because present TEFs are basedlargely on oral uptake studies often through diet. Application ofthese ‘‘intake or ingestion’’ TEFs for calculating the TEQ inabiotic environmental matrices has limited toxicological
relevance and use for risk assessment, unless the aspect of
reduced bioavailability and environmental fate and transport ofthe various dioxin-like compounds are taken into account. Ifhuman risk assessment is done for abiotic matrices, it isrecommended that congener-speciﬁc equations be usedthroughout the whole model, instead of using a total TEQbasis, because fate and transport properties differ widelybetween congeners.
A number of future approaches to determine alternative or
additional TEFs were identiﬁed. The use of a probabilisticmethodology to determine TEFs has the advantage that it betterdescribes the level of uncertainty in a TEF. The disadvantagecould be that this approach lumps data together and givessimilar weight to all studies, a problem that can only partly beavoided 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 regulatoryauthorities and international harmonization of TEFs. Further-more, 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 intakesituations. There is emerging experimental evidence which
suggests that some REPs may differ when based on adminis-tered dose versus tissue concentration. The development anduse of systemic TEFs and TEQ are recommended as anadditional approach to the present TEF concept, but at present,
there are insufﬁcient data to develop these systemic TEFs.
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