Toxicological Risk Assessment of Chemicals: A Practical Guide - Chapter 5 potx

5 Standard Setting: Threshold Effects Adve rse h ealth effect s can be consi dered to be of two types (see Secti on 4.2): those consi dered to have a threshold, k nown as ‘‘thres hold effects ’’ (effec ts such as, e.g., organ-speci fic, neurologica l, immunol ogical, non-gen otoxic carci nogenicity , reprod ucti ve, develo pment al), and those for whi ch there is consi dered to be some risk at any exposur e level, know n as ‘‘non-t hreshold effect s’’ (effects such as, e.g., mut agenicity, genotoxici ty, genoto xic carcinogeni city). Tho ugh it is not possi ble to demon strate experiment ally the presen ce or absence of a thres hold, diffe rences in the approac h to the hazard asses smen t of threshold versus non-thresho ld effects have been adopte d widely. The distinct ion in approac he s is based primari ly on the prem ise that simple events such as in vitro activati on an d covale nt bindi ng may be line ar over many orders of magnitud e, i.e., that these events occur even at very low exposur e levels. How ever, a simple pragm atic dist inction on this basis is incre asingly probl ematic as it is likely that there is a thres hold for a numbe r of genoto xic effect s; this is addres sed in detai l in Cha pter 6. In the hazard asses sment proces s, descri bed in detail in Cha pter 4, all effect s observ ed are evalua ted in terms of the type and severi ty (adver se or non-adv erse) , their dose –response relation- ship, and the relevanc e for human s of the effects observ ed in experi mental animals. For threshold effects, a No- or a Lowes t-Obser ved-Advers e-Effect Lev el (N=LOAE L), or alternati vely a Ben ch- mark Dos e (BMD), is deriv ed for every single effect in all the avail able studi es provi ded that data are suffi cient for such an evalua tion. In the last step of the hazard asses sment for thres hold effects, all this informat ion is asses sed in tota l in order to ident ify the critical effect(s) and to de rive a NOA EL, or LOAE L, for the critical effect(s) . The approac h of deriv ing a tole rable inta ke by divi ding the N=LOA EL, or alternati vely a BMD for the critical effect(s) by an asses smen t facto r has been descri bed and discu ssed extens ively in the scien tific literat ure. It is beyond the scope of this book to revie w all these refere nces. This chapte r presen ts an overvi ew of published extrapolat ion methods for the deriva tion of a tolerable inta ke based on the assessmen t factor approac h, i.e., limited to addres s effects with thres hold charact er- istics, and is not meant to be exhaust ive. The mai n focus is on the rationale for and the use of the asses sment facto rs. Pe rtinent guidan ce docum ents and revie ws for the issues add ressed in this chapte r include WHO=IPCS (1994, 1996, 1999), US-EPA (2002, 2004), IGHRC (2003) , ECETO C (2003) , KEMI (2003) , Kal berlah and Schneid er (1998) , Vermei re et al. (1999) , and Nielse n et al. (2005) . The approac h of standard setting for non-t hresho ld effect s is addres sed in Cha pter 6. The development of regulatory standards derived from a standard such as, e.g., the Tolerable Daily Intake or a Ref erence Dos e, is addres sed in Chapter 9. 5.1 INTRODUCTION Acco rding to the OECD=IPC S de finitions listed in Annexure 1 of Cha pter 1 (OECD 2003): Threshold is ‘‘Dose or exposure concentration of a substance below that a stated effect is not observed or expected to occur.’’ ß 2007 by Taylor & Francis Group, LLC. Tolerabl e Intak e is ‘‘Estimate d maxi mum amoun t of an agent, expres sed on a body mass basis, to which each indi vidual in a (sub) p opulation may be exp osed over a speci fied perio d wi thout appreci able risk .’’ A tolerable intake may have different units dependi ng on the route of adminis tration upon which it is based, an d is general ly expres sed on a dail y or weekly basis . For the oral and derm al route s, a tole rable intake is general ly expres sed on a body weight basis , e.g., mg=kg body wei ght per day. Tho ugh not strictly an ‘‘ intake, ’’ tolerable intakes for inhalati on are generally expres sed as an airborne concentration, e.g., mg=m 3 . Acco rding to the OEC D=IPCS de finitions listed in Anne xure 1 of Chapter 1 (OECD 2003): Acceptable=Tolerable Daily Intake is ‘‘Estimated maximum amount of an agent, expressed on a body mass basis, to which an individual in a (sub) population may be exposed daily over its lifetime without appreciable health risk.’’ Reference Dose is ‘‘An estimate of the daily exposure dose that is likely to be without deleterious effect even if continued exposure occurs over a lifetime.’’ Related terms: Acceptable=Tolerable Daily Intake. The term ‘‘acceptable’’ is used widely to describe ‘‘safe’’ levels of intake and is applied for chemicals to be used in food production such as, e.g., food additives, pesticides, and veterinary drugs. The term ‘‘tolerable’’ is applied for chemicals unavoidably present in a media such as contamin- ants in, e.g., drinking water and food. The term ‘‘PTWI’’ (Provisional Tolerable Weekly Intake) is generally used for contaminants that may accumulate in the body, and the weekly designation is used to stress the importance of limiting intake over a period of time for such substances. The tolerable intake is similar in definition and intent to terms such as ‘‘Reference Dose’’ and ‘‘Reference Concentration’’ (RfD=RfC), which are widely used by, e.g., the US-EPA. For some substances, notably pesticides, the ‘‘ARfD’’ (Acute Reference Dose), is also established, often from shorter-term studies than those that would support the ADI. The ARfD is defined as the amount of a substance in food that can be consumed in the course of a day or at a single meal with no adverse effects. In inhalation studies, laboratory animals are generally exposed to an airborne chemical for a limited period of time, e.g., 6 h a day, 5 days per week. Adjustment of such an intermittent exposure to a continuous exposure scenario is regularly applied as a default procedure to inhalation studies with repeated exposures but not to single-exposure inhalation toxicity studies. Operationally, this is accomplished by a correction for both the number of hours in a daily exposure period and the number of days per week that the exposures were performed. In an inhalation study in which animals were exposed to an airborne concent ration of a substance at 5 mg=m 3 for 6 h a day, for 5 days per week, the adjustment of this intermittent exposure concentration to a continuous exposure concentration would consider both hours per day and days per week : 5 mg=m 3 3 6=24 h 3 5=7 days=week ¼ 0.9 mg=m 3 , with 0.9 mg=m 3 being the concentration adjusted to continuous exposure. For systemi c effects observed in inhalation studies, the determining factor for effects to occur at the systemic target is generally the total dose rather than the concentration of the chemical in the air. In such cases, a tolerable intake (expressed as mg=kg body weight per day, or mg=m 3 depending on the standard to be derived, i.e., a tolerable intake in its strict meani ng, or a tolerable concentration) is established from the NOAEC, or LOAEC, derived in the inhalation study and adjusted for continuous exposure. For local effects, in contrast, the deter mining factor for effects to occur at the site of first contact (mucous membrane of the respiratory tract, the eyes, or the skin) is generally the concentration of the chemical in the air rather than the total dose at the site of first contact. In such cases, a tolerable concentration (expressed as mg=m 3 ) is established from the NOAEC, or LOAEC, derived in the inhalation study without an adjustment to a continuous exposure. The overall principles for the derivation of a tolerable intake are equal irrespective of chemical class (e.g., food additives, pesticides, veterinary drugs, contaminants) although it should be recog- nized that the available database for chemicals deliberately added to, e.g., food is generally more ß 2007 by Taylor & Francis Group, LLC. compr ehensive than for contam inants. This is because there a re extens ive regul atory demands for toxicit y data in relat ion to marketin g of che micals, which are intenti onally a pplied to food, etc. For threshold effect s, a tole rable intake is generally deriv ed from the NOA EL, or LOA EL, for the critical effect (s) by dividing the NOA EL=LOA EL, by an overall assessmen t facto r. Acco rding to the OECD=IPC S de finitions listed in Annex ure 1 of Chapter 1 (OECD 20 03): Assessm ent Fa ctor is ‘‘ Numeri cal adjus tment used to extrap olate from experi mentall y deter - mined (dose –respon se) relationsh ips to e stimate the agent exposur e below whic h an advers e effect is not like ly to occur. ’’ Relate d term s: Safety Factor , Unce rtainty Factor, Extrapo lation Factor , Adjust ment Factor , Conversi on Factor. The re is an enormous variability in the extent and natur e of different databa ses for chemical substanc es. Fo r examp le, in some cases, the evalua tion of a chemi cal must b e based on limit ed data in experi mental animals, whereas in other cases detai led infor mation on the vario us end- points, toxicokin etics, an d mode( s) of action may be available. In some cases, the evalua tion can be based on data on effect s in exposed human populatio ns. Clearl y as the amount of infor mation available incre ases, the degree of und erstanding of the hazards expres sed also incre ases, and the uncert ainties due to lack of informat ion decreas e. However , even with complex databa ses, uncer- tainties sti ll remain. The asses sment facto rs general ly applied in the estab lishmen t of a tolerable inta ke from the NOA EL, or LOA EL, for the crit ical effect (s) are appli ed in order to compe nsate for uncert ainties inher ent to extrapolat ion of experiment al animals data to a g iven human situation, and for uncer- tainties in the toxicolog ical databa se, i.e., in cases wher e the substanc e-speci fic knowledge requi red for risk asses smen t is not avail able. As a co nsequen ce of the variabili ty in the extent and natur e of different databa ses for chemi cal substances, the range of assessmen t facto rs applied in the estab - lishmen t of a tole rable intake has been wide (1 –10,000), althoug h a value of 100 has been used most often . An overview of diff erent approac hes in using assessmen t factors, historic ally and currently, is provi ded in Secti on 5.2. The key areas of uncert ainty when using data from experi mental anim als include uncertaint y related to: . Extrapo lation from anim al speci es to human s (Secti on 5.3) . Variabi lity in the human popula tion (Secti on 5.4) . Route-t o-route extra polation (S ection 5.5) . Durati on of exposur e in experi mental studies (Section 5.6) . Dose –respon se curve=NOA EL n ot estab lished (Section 5.7) . Nature and severity of the effects (Section 5.8) . Gaps or other de ficienci es in the databa se (Section 5.9) 5.2 ASSESSMENT FACTORS: GENERAL ASPECTS In the context of assessment factors, it is important to distinguish between the two terms ‘‘variabil- ity’’ and ‘‘uncertainty.’’ Variability refers to observed diff erences attributable to true heterogeneity or diversity, i.e., inherent biological differences between species, strains, and individuals. Variability is the result of natural random processes and is usually not reducible by further measurement or study although it can be better characterized. Uncertainty relates to lack of knowledge about, e.g., models, parameters, constants, data, etc., and can sometimes be minimized, reduced, or eliminated if additional information is obtained (US-EPA 2003). It should be recognized that a lack of knowledge of variability is a source of uncertainty. The terminology within this area is not standardized. Other terms include ‘‘safety factor,’’ ‘‘uncertainty factor,’’ ‘‘extrapolation factor,’’ ‘‘adjustment factor,’’ and ‘‘conversion factor.’’ None ß 2007 by Taylor & Francis Group, LLC. of these terms are ideal . For examp le, the term safety facto r has implic ations of absol ute safety, wher eas the term uncert ainty facto r, alth ough being broader, may be interpret ed different ly in relation to varia bility a nd uncert ainty. For the sake of clarity in this book, the term asses sment facto r is used and is meant as a g eneral term to cover all facto rs desig nated in the literatu re as safety facto r, unce rtainty facto r, extrapolat ion facto r, adjus tment facto r, convers ion facto r, etc. The other ment ioned term s are not used unless refere nce is made to a speci fic term or met hod. The asses sment facto r can cover both varia bility and uncert ainty. The follow ing secti on gives an overview of different approac hes in using assessmen t factors, histori cally and curren tly, beginn ing with the introduct ion of the so-cal led ‘‘safet y factor approac h’’ in the mid-1950s and re flecting the develo pment up to the regul atory approac hes c urrently used by inte rnational and federal b odies. The overview does not attempt to cover all publi cations in this field, but includes the approac hes sugges ted by diff erent scien tifi c groups and inte rnational and federal bodies , which are considered as being the most central ones in the development of the approaches currently used regulatory. Default assessment factors used or suggested in the various approac hes are summ arized in Table 5 .1. 5.2.1 ASSESSMENT FACTORS:VARIOUS APPROACHES Historically, the so-called safety factor approach was introduced in the United States in the mid- 1950s in response to the legislative needs in the area of the safety of chemical food additives (Lehman and Fitzhugh 1954). This approach proposed that a ‘‘safe level’’ of chemical food additives could be deriv ed from a chronic NOAEL from animal studies divided by a 100-fold safety factor. The 100-fold safety factor as proposed by Lehman and Fitzhugh was based on a limited analysis of subchronic=chronic data on fluorine and arsenic in rats, dogs, and humans, and also on the assumption that the human population as a whole is heterogeneous. Initially, Lehman and Fitzhugh reasoned that the safety factor of 100 accounted for several areas of uncertainty: . Intraspecies (human-to-human) variability . Interspecies (animal-to-human) variability . Allowance for sensitive human populations due to illness when compared with healthy experimental animals . Possible synergistic action of the many intentional and unintentional food additives or contaminants In 1961, the Joint FAO=WHO Expert Committee on Food Additives (JECFA) and the Joint Meeting of Experts on Pesticides Residues (JMPR) adopted this approach in a slightly modified form: The safe level was called the Acceptable Daily Intake (ADI) and expres sed in mg=kg body weight per day (Vermeire et al. 1999, ECETOC 2003). Usually, a safety facto r of 100 is used by JECFA and JMPR for establishing ADIs by this ADI approach; however, the procedures adopted by JECFA and JMPR do not generate a clear justification for deviation from the factor of 100, but in some individual cases, an expert explanation is given for the use of factors other than 100 (Vermeire et al. 1999). It is apparent that the factor of 100 has no quantitative bases, and the choice of the value 100 is more or less arbitrary (Vermeire et al. 1999). Retrospectively, some attempts have been made to support a 100-fold factor (Bigwood 1973, Lu 1979, Vettorazzi 1977 as reviewed in Vermeire et al. 1999 and KEMI 2003), and the 100-fold factor was found to be justified. The 100-fold safety factor has tradi tionally been interpreted as the product of two factors with default values of 10. For example, according to WHO=IPCS (1987), the safety factor is intended to provide an adequate Margin of Safety (MOS) by assuming that the human being is 10 times more sensitive than the test animal and that the difference of sensitivity within the human population is in a 10-fold range. ß 2007 by Taylor & Francis Group, LLC. TABLE 5.1 Default Assessment Factors Used or Suggested for the Establishment of a Regulator y Standard or Health-Base d Guidance Value for Threshold Effects Factor JECFA JMPR US-EPA (2002) Renwick (1993) WHO LLN (1990) ECETOC (2003) TNO (1996) Kalberlah and Schneider (1998) KEMI (2003) a D-EPA (2006) Interspecies 10 10 10 10 Toxicokinetic A b 44 A b 4 (rat) Toxicodynamic 10 0.5 2.5 2.5 1 2.5 Oral A b 3 3A b 3 2– 3 Inhalation 3 Interindividual 10 10 10 5 10 25 10 Toxicokinetic 4 3.16 8 3–5 Toxicodynamic 2.5 3.16 3 3.16 Occupational 33 Route-to-route Duration exposure 10 Subacute-to-subchronic 10 3 Subacute-to-chronic 624 Subchronic-to-chronic 10 2 10 8 10 LOAEL-to-NOAEL 10 10 3 1 3–10 10 Nature and severity 11–10 up to 10 Confidence database 10 10 1 1–10 1–10 Nonscientific1 a See also Table 5.2. b A is a calculated adjustment factor allowing for the differences in caloric requirement. ß 2007 by Taylor & Francis Group, LLC. 5.2.1. 1 US-E PA Appr oach In 1988, the US-EPA adopte d the ADI approac h in its regul atory meas ures against envir onmen tal poll ution; with a numbe r of modi fi cations (US-E PA 1988, 1993). Inst ead of the terms ADI and safet y facto r, the term s Reference Dose (RfD) and uncert ainty facto r (UF), respec tively, wer e selec ted. The RfD is deriv ed from the NOA EL by divi ding by the overall UF. The o verall UF origina lly sugges ted and recon firmed in 2002 (US-E PA 2002) general ly consi sts of a 10-fol d factor for each of the following: . Huma n varia tion in sensi tivity (UF H ) . Inter species extra polation (UF A ) . Use of the NO AEL obtained from a less than lif etime study (UF S ) . Use of a LOAE L in the absence of a NOAEL (UF L ) . Adeq uacy of the total database (UF D ) Acco rding to US-EPA (1993), the fi rst four of the above-ment ioned factors a re adapte d from Dour son and Stara (1983). The exact value of the UFs chosen shoul d depend on the quali ty of the studie s av ailable, the extent of the database, and scien tific judgment (US- EPA 2002 ). The default facto rs typic ally used cover a single order of magni tude (i.e., 10 1 ). By convent ion, in the US-EPA , a value of 3 is used in place of one-hal f powe r (i.e., 10 0.5 ) when appropriat e. The se half- power values should be facto red as whole numbe rs when they occur singly but as powers or logs when they occur in tandem . A composit e UF of 3 and 10 would thus be expres sed as 30 (3 3 10 1 ), whereas a compo site UF of 3 and 3 would be expressed as 10 (10 0.5 3 10 0.5 ¼ 10 1 ). It shoul d be n oted, in addition, that rigid applicati on of log (i.e., 10 1 ) or ½ log (i.e., 10 0.5 ) un its for UFs could lead to an illogic al set of refere nce values ; therefore, it ha s been empha sized that appli cation of scien tific judgm ent is critical to the overall proces s. It is also noted that there is overl ap in the individua l UFs and that the appli cation of five UFs of ten for the chronic reference value (yielding a total UF of 100,000) is inappr opria te. In fact, in cases wher e maxi mum uncertaint y ex ists in a ll fi ve areas, it is unlikely that the databa se is suf ficient to deriv e a refere nce value. Uncertainty in four areas may also indi cate that the database is insuf fi cient to deriv e a reference value. In the c ase of the RfC, the maxi mum UF would be 3,000, whereas the maxi mum would be 10,000 for the RfD. Thi s is because the deriv ation of RfCs and RfDs has evolve d some what different ly. The RfC methodol ogy (US- EPA 1994) recommend s divi ding the inte rspecies UF in half, one-half (10 0.5 ) each for toxicokin etic and toxi codynam ic consi derations, and it includes a Dosimetric Adjustment Factor (DAF, represents a multiplicative factor used to adjust an observed exposure concentration in a particula r laboratory species to an exposure concentration for humans that would be associated with the same delivered dose) to account for toxicokinetic differences in calculating the Human Equivalent Concentration (HEC), thus reducing the interspecies UF to 3 for toxicodynamic issues. RfDs, however, do not incorporate a DAF for deriving a Human Equ ivalent Dos e (HED), and the interspeci es UF of 10 is typi cally applied, see also Secti on 5.3 .4. It is recommended to limit the total UF applied for any particular chemical to no more than 3000, for both RfDs and RfCs, and avoiding the derivation of a reference value that involves application of the full 10-fold UF in four or more areas of extrapolation. In addition, a modifying factor (MF) could be applied (US-EPA 1993). The MF is in reality an additional UF that is greater than 0 and less than or equal to 10; the default value is 1. The MF should account for uncertainties of the study and database not explicitly handled by the use of the general UFs; e.g., the completeness of the overall database and the number of species tested. In the 2002 review of the RfD and RfC processes (US-EPA 2002), it was recommended that use of the MF be discontinued as it was considered that the uncertainties accounted for by the MF is sufficiently accounted for by the general UF. The US-EPA staff paper from 2004 titled ‘‘An Examination of EPA Risk Assessment Principles and Practices’’ (US-EPA 2004) provides comprehensive and detailed information on the ß 2007 by Taylor & Francis Group, LLC. pract ices empl oyed in risk asses smen t, incl uding use of UFs and use of default and extra polation assumpti ons. 5.2.1. 2 Calabres e and Gilbert Appro ach Calabres e and Gilber t (1993) hav e demonstra ted the lack of indepe ndence of the interspeci es and intraspeci es UFs, as well as of the intraspeci es and the less -than-lifeti me UFs. Based on thei r analys es, the author s conclu ded that most of the recommend ed US-EPA stand ards based on animal models needed to have some of their UFs modi fied. They recom mended the following modi fi cations of the intraspeci es UF, see also Secti on 5.4 .2: . Signi ficantly less-than -lifeti me anim al study : 5 . When the anim al study is for a norm al experi mental lifetime (2 years in rodent s): 4 . Occu pational epidemiolo gical study: 10 . Environ menta l epidemiol ogical study , if study was for a no rmal human life span: 5 5.2.1. 3 Renwic k Appro ach The approac h propos ed by Renwick (1991, 1993) is also b ased on the 100-fo ld facto r. It attempt s to give a scien tifi c basis to the default values of 10 for the interspeci es and 10 for the intraspeci es (interin dividual human ) diff erences. Ren wick also propos ed a divi sion of each of these UFs into sub-f actors to allo w for separa te evalua tions of diff erences in toxicokin etics and toxicodyna mics. The advanta ge of such a subdi vision is that compo nents of these UFs can be addres sed wher e data are avail able; for example, if ava ilable data show simil ar toxicokin etic s of a given chemical in experi mental anim als and humans, then only an inte rspecies e xtrapolat ion facto r would be nee ded to account for diff erences in toxi codynam ics. Renwi ck examined the relative magnitud e of toxicoki- netic and toxicodynam ic varia tions between and wi thin species in detail. He found that toxi cokinetic differences were generally greater than toxicodyn amic d ifferences resul ting in the propos al that the 10-fold facto rs (for inter- and intraspeci es variation ) shoul d, by defaul t, be subdi vided into factors of 4 for toxicokin etics and 2.5 for toxi codynam ics. It should be no ted that the propos ed defaul t values were deriv ed from limited data. The WHO=IPCS (1994) ha s adopte d the approac h set forth by Ren wick (1993) with one deviation, see Figure 5.1. Whil e the UF for inte rspecies (animal- to-human) extra polation shoul d be split into defaul t values of 4 for toxi cokinetics and 2.5 for toxicodynam ics, the UF for intraspeci es (hum an-to-hum an) extra polation shoul d be split evenly be tween both aspect s, i.e., a sub-f actor of 3.16 for both toxicokin etics and toxicodynam ics. The reason for this deviation from Renwi ck ’s initial sugges tion was that the WHO=IPCS consi dered that the slightly greater varia bility in the kinetics in human s compa red with dy namics was not suf ficient to warrant an unequal subdivisi on of the 10-fol d facto r into a toxic okinetic facto r of 4 and a toxicodynam ic facto r of 2.5. Actual data shoul d be used to repla ce the defaul t values if availab le. It was furt hermore noted that precise defaul t values for kinetics and dy namics cannot be expect ed on the basis of a subdivisio n of the imp recise 10-fold compo site facto r for interspeci es as wel l as for the inte rindivid ual varia tion. Acco rding to WHO=IPCS, the defaul t v alues sugges ted above wer e consi dered as being reason able since they provide a positive value greater than 2 for both aspects and are compatible with the species differences in physiological parameters such as renal and hepatic blood flow. It was also noted that since the database examined was limited, the default values suggested for subdivision of interspecies and interindividual variation should be adopted on an interim basis. 5.2.1. 4 Lewis –Lyn ch–Nikiforov Appr oach In 1990, Lewis et al. published a new approach introducing flexibility such that both new information and expert judgment could be readily incorporated. The Lewis–Lynch–Nikiforov (LLN) method, and its refinements, are extensions of established principles and procedures, and ß 2007 by Taylor & Francis Group, LLC. guides the data evalua tor to adjus t experi mentall y determin ed ‘‘no-eff ect ’’ (or ‘‘ minim um effect ’’) level s from expe rimental animal studies taking the foll owing aspects into account : . Known diff erences between labor atory anim als and human s and between experi mental condit ions and the real world . Sensit ivity of the exposed human popula tions . Streng th of eviden ce that the chemi cal presen ts a real haz ard to human healt h . Gene ral quali ty of the experi mental databa se . Unce rtaintie s in extrapolat ing from labor atory anim als to h umans . Potency of the toxic agent . Typ e and severi ty of the putat ive advers e effect Acco rding to Lewis et al. (1990) , a step- by-step sequenc e is used. Initially, a quali tative deter min- atio n is made as to the strength o f eviden ce that the putative toxi c agent presen ts an a ctual health hazard to human s, i.e., how like ly is this agent to p roduce the suspec ted adverse effect in human s? In contrast to the ADI and RfD method wher e no speci fic c onsideration is given to judgi ng the like lihood that a chemical presents a real health hazard, the ‘‘stre ngth of the quali tative evidence ’’ is scored expli citly and separa tely in the LLN approac h. The NA EL human is estimat ed from laboratory resear ch results, using the follow ing a lgorithm: NAE L human ¼ NOA EL animal [S] [I][R][ Q 1 ][Q 2 ][Q 3 ][U][ C] [S] is the aggrega te ‘‘scali ng facto r ’’ to account for known quanti tative diff erences betw een species and between labor atory experi mental condition s and the real world. The default value is 1, indi cating that animals and humans are equivalent in these dim ension s. [I] is the adjus tme nt facto r to account for anticipat ed g reater suscep tibility among mem bers of the test animal populatio n than was ob served in the experi ment, i.e., to account for intraspeci es varia bility. The default value is 10, indi cating that extre mely high variability was observ ed (or would be expect ed) a mong anim als. Uncertainty factor 100 Interspecies differences (differences between humans and common laboratory animals) 10 Toxico- dynamics 10 0.4 (2.5) Intraspecies differences (differences among humans; interindividual differences) 10 Toxico- kinetics 10 0.6 (4.0) Toxico- dynamics 10 0.5 (3.2) Toxico- kinetics 10 0.5 (3.2) FIGURE 5.1 Subdivision of the 100-fold UF showing the relationship between the use of UFs (above the dashed line), and the proposed subdivisions (below the dashed line) based on toxicokinetics and toxicody- namics. (From Renwick, A.G., Food Addit. Contam., 10, 275, 1993; WHO=IPCS Assessing human health risks of chemicals: Derivation of guidance values for health-based exposure limits. Environmental Health Criteria 170. Geneva, 1994. Available at http:== www.inchem.org=documents=ehc=ehc=ehc170.htm) ß 2007 by Taylor & Francis Group, LLC. [R] is the adjus tment facto r to account for anti cipated differences in suscep tibility be tween human s and the laboratory animals, i.e., to account for inte rspecies varia bility. The defaul t value is 10, indicati ng that humans are much more susceptible. [Q 1–3 ] and [U] are adjustm ent factors to account for varia tions in the reliabilit y of the databa se (data quali ty) and other source s of uncert ainty in the data e valuation p rocess. [Q 1 ]reflects the data evalua tor’ s certa inty that the agent actually causes the speci fic ‘‘critical effect ’’ in human s. The default value is 1, indi cating that the agent causes similar toxic effect s in animals and humans. [Q 2 ] is employed when e xtrapolat ing data from subchro nic studies to estimat e risk from lifelong exposur es. The defaul t value 10, indi cating great uncert ainty in esti mating the NOAEL chronic from the NO AEL subchronic . [Q 3 ] is empl oyed when e xtrapolat ing LOAE Ls to NOA ELs. The defaul t value 10, indicati ng extre mely great uncert ainty associated with using a LOAE L animal to estimate a NAEL human . [U] is used to account for resi dual uncertaint y in estimates of [S], [I], and [R]. The default value is 10 indicati ng very great overal l uncertaint y, which has not alrea dy been account ed for in [Q 1–3 ]. [C] is a nonsci enti fi c, judgm ental safet y factor, i.e., a socia l or poli tical value judgm ent. The defaul t value is 1, indicating that no addit ional MOS is needed over that provi ded by the inheren tly conser vative procedu re above. An aggrega te adjus tment factor of about 2 50 is typical; the theor etical maximum v alue is 100,000. By appli cation of factors [Q 1–3 ] and [U], this approac h attempted to separa te scien tific judg- ments from poli cy=value judgmen ts. Acco rding to the author s, there are three distingu ishing featu res of the LLN approac h. The first is the empha sis on careful discr iminat ion among the adjus tments. The second is on discr iminat ion betw een ‘‘best esti mates ’’ of the correct adjus tments for [S], [I], and [R] and the compl etely separa te adjus tment for overal l un certainty . The thir d is on securing scien tific consens us on the adjus tment values . It should be recognized, howe ver, that in pract ice, it will not be possible to dist inguish all these facto rs, and that some factors may not be indepe ndent of each other. It could also be que stioned whether a nonsci enti fic factor [C] shoul d be discussed in a scien tific risk asses sment. 5.2.1.5 EU TGD Approach The process of human health risk assessment has been extensively addressed within the EU framework of Risk Assessment of New and Existing Chemical Substances. According to the EU Technical Guidance Document (TGD) on Risk Assessment of New and Existing Chemical Substances (EC 1996), the risk characterization is carried out by quantitatively comparing the outcome of the effects assessment to the outcome of the exposure assessment, i.e., a comparison of the NO AEL, or LOAE L, and the exposur e estimate, see Secti on 8.3.3. The ratio resul ting from this comparison is called the Margin of Safety (MOS). The TGD recommends the following parameters to be considered in assessing the MOS: . Uncertainty arising, among other factors, from the variability in the experimental data and intra- and interspecies variation . Nature and severity of the effect . Human population to which the quantitative and=or qualitative information on exposure applies . Differences in exposure (route, duration, frequency, and pattern) . Dose–response relationship observed . Overall con fidence in the database These parameters are parallel to those being considered in the evaluation of the assessmen t factors to be applied in the establishment of a tolerable intake. ß 2007 by Taylor & Francis Group, LLC. The TGD has been revised and the second edition was published in 2003 (EC 2003). However, the human health risk characterization part was not included in this second edition. A final draft version of the human health risk characterization part was released in 2005 with a d etailed guidance on, among other s, the main issues to be included in derivation of the ‘‘reference MOS’’ (MOSref), which is analogous to an overall assessment factor. The individual factors contributing to the MOSref are described separately and guidance is given on how to combine these into the MOSref. The guidance provided in this draft version has been extensively used in relation to the risk assessment of prioritized substances carried out since the draft version was released; however, this version is not publicly available. In the new EU chemicals regulation REACH, which entered into force on 1 June 2007, detailed guidance documents on different REACH elements, including risk characterization and the use of assessment factors, are currently in preparation (spring 2007). These documents will probably be available on the EU DG Environment REACH Web site (EU 2006) when published. 5.2.1.6 ECETOC Approach The approach recommended by the ECETOC (1995) is to derive the best scientific estimate of a Human No-Adverse-Effect Level, referred to in the report as the Predicted No-Adverse-Effect Level (PNAEL). The approach distinguishes three stages: . Application of a scientifically derived adjustment factor to the NOAEL, or LOAEL, of the critical effect established in the pivotal study. It is stated that if the database is inadequate, then human PNAELs cannot be derived scientifically. . Application of a UF to the PNAEL to take into account the degree of scientific uncertainty involved. The following degrees of confidence in the human PNAEL are suggested: high ¼ 1, medium ¼ 1–2, low ¼ larger UF. . Application of a nonscientifically based safet y factor to take into account political aspects, socioeconomic aspects (cost–bene fit considerations), or risk perception factors (the nature of the effect may justify the use of an additional factor). The scientifically derived adjustment factors include the following elements: . Experimental exposure in relation to the expected human exposure: a default value of 3 for extrapolation from short-term to subchronic exposure; a default value of 2– 3 for extrapol- ation from subchronic to chronic exposure . Extrapolation from LOAEL to NOAEL: a default value of 3 . Route-to-route extrapolation: no default value . Interspecies extrapolation (animal-to-human): a default value of 4 for oral exposure (for the rat with a body weight of 250 g and based on caloric demands); a default value of 1 for inhalation . Intraspecies extrapolation (human-to-human): a default value of 3 for the general popula- tion; a default value of 2 for workers This approach discriminates factors to a large extent in order to distinguish between the single adjustments and to separate best estimates from uncertainty. It should be noted that the ECETOC approach does not mention the establishment of an overall factor and although they mention that all discriminated aspects introduce uncertainties, they do not give guidance on how to account for this. It could also be questioned here whether a nonscientific factor should be discussed in a scientific risk assessment. In a more recent report, ECETOC (2003) has further developed many of the principles established in the previous report (ECETOC 1995) and replaced the guidance provided there in on ß 2007 by Taylor & Francis Group, LLC. [...]... conclusions drawn from the evaluation of the available data on default assessment factors was that the conventionally used factor of 100 (10 for animal-to-human and 10 for human-to-human variations) is probably an underestimate It is stated that it is likely that the animal-to-human extrapolation is greatly underestimated, and in the case of human-to-human variability, an assessment factor of 10–16 is... (animal-to-human) factor and the subchronic-to-chronic duration factor are considerably higher than 10 In addition, the limited data on intraspecies (human-to-human) variation is also considered to indicate that a default factor of 10 may not be sufficient Derivation of approximations of the distribution of assessment factors from historical data (based on NOAEL ratios) has limitations as the use of. .. information include: LOAEL-to-NOAEL extrapolation: a default value of 3 Duration extrapolation: subacute to chronic, a default value of 6; subchronic to chronic, a default value of 2; local effects by inhalation, a default value of 1 Route-to-route extrapolation: oral to inhalation, no default value; oral to dermal, no default value Interspecies extrapolation (animal-to-human): systemic effects (scaling):... mouse, a default value of 7; rat, a default value of 4; monkey, a default value of 2; dog, a default value of 2 Local effects by inhalation: a default value of 1 Intraspecies extrapolation (human-to-human): systemic and local effects: a default value of 5 for the general population; a default value of 3 for workers Similarly to the previous ECETOC approach, this revised approach does not mention the establishment... means an extrapolation factor of 7 for mouse-to-man, 3.9 for rat(Fischer) -to-man, 3.6 for rat(Sprague–Dawley)-to-man, 1.6 for dog(Beagle)-to-man, 3.9 for monkey (marmoset)-to-man, and 1.6 for monkey(rhesus)-to-man It was noted that these extrapolation factors only account for toxicokinetic differences in the basal metabolic rate If an interspecies extrapolation with an average degree of statistical... a default value of 10; other aspects, a default value of 1 Type of critical effect: a default value of 1 Dose–response curve: a default value of 1 Confidence of the database: a default value of 1 Route-to-route: no default value Principally, the overall assessment factor is established by multiplication of the separate factors The authors note that in practice it is not possible to distinguish all above-mentioned... the Environment) and TNO The report concentrated on the quantification of default distributions of the assessment factors related to interspecies extrapolation (animal-to-human), intraspecies extrapolation (human-to-human), and exposure duration extrapolation 5. 2.1.8 Kalberlah and Schneider Approach In a report on a research project ‘‘quantification of extrapolation factors’’ (Kalberlah and Schneider 1998),... Vermeire et al (1999, 2001) and KEMI (2003), or a deterministic default factor of 2 .5 could be used for extrapolation of data from rat studies to the human situation ß 2007 by Taylor & Francis Group, LLC 5. 4 INTRASPECIES EXTRAPOLATION (INTERINDIVIDUAL, HUMAN-TO-HUMAN) Risk assessments are usually based on data from studies in animals of similar age In addition, the animals are initially healthy and are fed... in an experimental animal study, Xanimal, and the equivalent dose in man, Xhuman, the scaling factor for man from the experimental animal is Scaling ¼ Xhuman =Xanimal ¼ a( Whuman )n =a( Wanimal )n and Xhuman ¼ Xanimal  [Whuman =Wanimal ]n To correct for differences in body size between humans and experimental animals, three measures of body size are used in practice as the basis for the extrapolation:... situation and for uncertainties in the toxicological database The assessment factors should be derived considering the toxicity profile of the substance; if the available data are insufficient, an overall assessment factor is used comprising various sub-factors related to: Interspecies differences (animal-to-human): mouse, a default value of 7 3 3; rat, a default value of 4 3 3; rabbit, a default value . that the animal-to-human extrapolation is greatly underestimated, and in the case of human-to-human variability, an assessment factor of 10–16 is considered as a minimum. Attention is also drawn. Human Health Risk Assessment Area to be Extrapolated Assessment Factor Deterministic Approach Probabilistic Approach Adequacy of the toxicological database relevance, validity, reliability 1 5. Initially, Lehman and Fitzhugh reasoned that the safety factor of 100 accounted for several areas of uncertainty: . Intraspecies (human-to-human) variability . Interspecies (animal-to-human) variability . Allowance