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# Reference Dose (RfD): Description and Use in Health Risk Assessments Background Document 1A March 15, 1993
Table of Contents 1.1. INTRODUCTION 1.1.1. BACKGROUND AND SUMMARY 1.1.2. OVERVIEW 1.2. TRADITIONAL APPROACH TO ASSESSING SYSTEMIC TOX ICITY 1.2.1. DESCRIPTION OF THE TRADITIONAL APPROACH 1.2.2. SOME DIFFICULTIES IN UTILIZING THE TRADITION AL APPROACH 1.3. EPA ASSESSMENT OF RISKS ASSOCIATED WITH SYSTEM IC TOXICITY 1.3.1. HAZARD IDENTIFICATION 1.3.2. DOSE-RESPONSE ASSESSMENT 1.3.3. EXPOSURE ASSESSMENT 1.3.4. RISK CHARACTERIZATION 1.4. APPLICATION IN RISK MANAGEMENT 1.5. OTHER DIRECTIONS 1.6. HYPOTHETICAL, SIMPLIFIED EXAMPLE OF DETERMININ G AND USING RfD 1.6.1. EXPERIMENTAL RESULTS 1.6.2. ANALYSIS 1.7. REFERENCES
1. REFERENCE DOSE (RfD): DESCRIPTION AND USE IN HEA LTH RISK ASSESSMENTS 1.1. INTRODUCTION This concept paper describes the U.S. Environmental Protection Agency's (U.S. EPA) principal approach to and rationale for assess ing risk for health effects other than cancer and gene mutations from chronic chemica l exposure. By outlining principles and concepts that guide EPA risk assessm ent for such systemic effects the paper complements the new risk assessment guide lines (U.S. EPA, 1987), which describe the Agency's approach to risk assess ment in other areas, specifically carcinogenicity, mutagenicity, develop mental toxicity, exposure, and chemical mixtures. (In this document the term "syst emic toxicity" refers to an effect other than carcinogenicity or mutagenicity i nduced by a toxic chemical.) 1.1.1. BACKGROUND AND SUMMARY Chemicals that give rise to toxic endpoints other t han cancer and gene mutations are often referred to as "systemic toxicants" becau se of their effects on the function of various organ systems. In addition, che micals that cause cancer and 

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12/5/2011 http://www.epa.gov/iris/rfd.htmgene mutations also commonly evoke other toxic effe cts (i.e., systemic toxicity). Based on our understanding of homeostatic and adapt ive mechanisms, systemic toxicity is treated as if there is an identifiable exposure threshold (both for the individual and for populations) below which there a re no observable adverse effects. This characteristic distinguishes systemic endpoints from carcinogenic and mutagenic endpoints, which are often treated as nonthreshold processes. Systemic effects have traditionally been evaluated using such terms as "acceptable daily intake (ADI)," "safety factor (SF )," and "margin of safety (MOS)," concepts that are associated with certain l imitations described below. The U.S. EPA established the Reference Dose (RfD) W ork Group to address these concerns. In preparing this report, the RfD Work Group has dr awn on traditional report on risk assessment (NRC, 1983), to more fully articula te the use of noncancer, nonmutagenic experimental data in reaching regulato ry decisions about the significance of exposures to chemicals. In the proc ess, the Work Group has coined less value-laden terminology -- "reference d ose (RfD)," "uncertainty factor (UF)"; "margin of exposure (MOE)"; and "regulatory dose (RgD)" -- to clarify and distinguish between aspects of risk assessment and risk management. These concepts are currently in general use in many parts of U.S. EPA. 1.1.2. OVERVIEW This document consists of four parts in addition to this Introduction. In Section 1.2., the traditional approach to assessing risks o f systemic toxicity is presented, and issues associated with this approach are identi fied and discussed. In Section 1.3., the modifications made to the traditional app roach by the Work Group are presented. Section 1.4. examines how these new conc epts can be applied in reaching risk management decisions, and Section 1.5 . briefly discusses some of the additional approaches the U.S. EPA is using and exploring to address this issue. Section 1.6. provides a sample RfD calculati on. Section 1.7. consists of references. 1.2. TRADITIONAL APPROACH TO ASSESSING SYSTEMIC TOX ICITY The U.S. EPA's approach to assessing the risks asso ciated with systemic toxicity is different from its approach to assessing the ris ks associated with carcinogenicity, because of the different mechanism s of action thought to be involved in the two cases. In the case of carcinoge ns, the Agency assumes that a small number of molecular events can evoke changes in a single cell that can lead to uncontrolled cellular proliferation. This m echanism for carcinogenesis is referred to as "nonthreshold," since there is theor etically no level of exposure for such a chemical that does not pose a small, but fin ite, probability of generating a carcinogenic response. In the case of systemic toxi city, however, organic homeostatic, compensating, and adaptive mechanisms exist that must be overcome before a toxic endpoint is manifested. For example, there could be a large number of cells performing the same or simila r function whose population must be significantly depleted before the effect is seen. The threshold concept is important in the regulator y context. The individual threshold hypothesis holds that a range of exposure s from zero to some finite 

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12/5/2011 http://www.epa.gov/iris/rfd.htmvalue can be tolerated by the organism with essenti ally no chance of expression

of the toxic effect. Further, it is often prudent t o focus on the most sensitive members of the population; therefore, regulatory ef forts are generally made to keep exposures below the population threshold, whic h is defined as the lowest of the thresholds of the individuals within a populati on. 1.2.1. DESCRIPTION OF THE TRADITIONAL APPROACH In many cases, risk decisions on systemic toxicity have been made by the Agency using the concept of the "acceptable daily intake ( ADI)" derived from an experimentally determined "no-observed-adverse-effe ct level (NOAEL)." The ADI is commonly defined as the amount of a chemical to which a person can be exposed on a daily basis over an extended period of time (usually a lifetime) without suffering a deleterious effect. The ADI con cept has often been used as a tool in reaching risk management decisions (e.g., e stablishing allowable levels of contaminants in foodstuffs and water.) A NOAEL is an experimentally determined dose at whi ch there was no statistically or biologically significant indication of the toxic effect of concern. In an experiment with several NOAELs, the regulatory focu s is normally on the highest one, leading to the common usage of the term NOAEL as the highest experimentally determined dose without a statistica lly or biologically significant adverse effect. The NOAEL for the critical toxic ef fect is sometimes referred to simply as the NOEL. This usage, however, invites am biguity in that there may be observable effects that are not of toxicological si gnificance (i.e., they are not "adverse"). For the sake of precision, this documen t uses the term NOAEL to mean the highest NOAEL in an experiment. In cases i n which a NOAEL has not been demonstrated experimentally, the term "lowest- observed-adverse-effect level (LOAEL)" is used. Once the critical study demonstrating the toxic eff ect of concern has been identified, the selection of the NOAEL results from an objective examination of the data available on the chemical in question. The ADI is then derived by dividing the appropriate NOAEL by a safety factor ( SF), as follows: ADI (human dose) = NOAEL (experimental dose)/SF. (E quation 1) Generally, the SF consists of multiples of 10, each factor representing a specific area of uncertainty inherent in the available data. For example, a factor of 10 may be introduced to account for the possible diffe rences in responsiveness between humans and animals in prolonged exposure st udies. A second factor of 10 may be used to account for variation in suscepti bility among individuals in the human population. The resultant SF of 100 has been judged to be appropriate for many chemicals. For other chemicals, with data base s that are less complete (for example, those for which only the results of subchr onic studies are available), an additional factor of 10 (leading to a SF of 1000) m ight be judged to be more appropriate. For certain other chemicals, based on well-characterized responses in sensitive humans (as in the effect of fluoride o n human teeth), an SF as small as 1 might be selected. While the original selection of SFs appears to have been rather arbitrary (Lehman and Fitzhugh, 1954), subsequent analysis of data (D ourson and Stara, 1983) 

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12/5/2011 http://www.epa.gov/iris/rfd.htmlends theoretical (and in some instances experiment al) support for their

selection. Further, some scientists, but not all, w ithin the EPA interpret the absence of widespread effects in the exposed human populations as evidence of the adequacy of the SFs traditionally employed. 1.2.2. SOME DIFFICULTIES IN UTILIZING THE TRADITION AL APPROACH 1.2.2.1. Scientific Issues While the traditional approach has performed well o ver the years and the Agency has sought to be consistent in its application, obs ervers have identfied scientific shortcomings of the approach. Examples include the following: a. Too narrow a focus on the NOAEL means that infor mation on the shape of the dose-response curve is ignored. Such data could be important in estimating levels of concern for public safety. b. As scientific knowledge increases and the correl ation of precursor effects (e.g., enzyme induction) with toxicity becomes known, ques tions about the selection of the appropriate "adverse effect" arise. c. Guidelines have not been developed to take into account the fact that some studies have used larger (smaller) numbers of anima ls and, hence, are generally more (less) reliable than other studies. These and other "scientific issues" are not suscept ible to immediate resolution, since the data base needed is not yet sufficiently developed or analyzed. U.S. EPA work groups are presently considering these iss ues. 1.2.2.2. Management-related Issues 1.2.2.2.1. The use of the term "safety factor" The term "safety factor" suggests, perhaps inadvert ently, the notion of absolute safety (i.e., absence of risk). While there is a co nceptual basis for believing in the existence of a threshold and "absolute safety" asso ciated with certain chemicals, in the majority of cases a firm experimental basis for this notion does not exist. 1.2.2.2.2. The implication that any exposure in exc ess of the ADI is "unacceptable" and that any exposure less than the ADI is "acceptable" or " safe" In practice, the ADI is viewed by many (including r isk managers) as an "acceptable" level of exposure, and, by inference, any exposure greater than the ADI is seen as "unacceptable." This strict demarcat ion between what is "acceptable" and what is "unacceptable" is contrary to the views of most toxicologists, who typically interpret the ADI as a relatively crude estimate of a level of chronic exposure which is not likely to re sult in adverse effects to humans. The ADI is generally viewed by risk assesso rs as a "soft" estimate, whose bounds of uncertainty can span an order of ma gnitude. That is, within reasonable limits, while exposures somewhat higher than the ADI are associated with increased probability of adverse effects, that probability is not a certainty. 

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12/5/2011 http://www.epa.gov/iris/rfd.htmSimilarly, while the ADI is seen as a level at whic h the probability of adverse

effects is low, the absence of all risk to all peop le cannot be assured at this level. 1.2.2.2.3. Possible limitations imposed on risk man agement decisions Awareness of the "softness" of the ADI estimate, as discussed above, argues for careful case-by-case consideration of the toxicolog ical implications of individual situation, so that ADIs are not given a degree of s ignificance that is scientifically unwarranted. In addition, the ADI is only one facto r in a risk management decision and should not be used to the exclusion of other relevant factors. 1.2.2.2.4. Development of different ADIs by differe nt programs In addition to occasionally selecting different cri tical toxic effects, Agency scientists have reflected their best scientific jud gments in the final ADI by adopting factors different from the standard factor s listed in Table 1. For example, if the toxic endpoint for a chemical in ex perimental animals is the same as that which has been established for a related ch emical in humans at similar doses, one could argue for an SF of less than the t raditional 100. On the other hand, if the total toxicologic data base is incompl ete, one could argue that an additional SF should be included, both as a matter of prudent public policy and as an incentive to others to generate the appropriate data.
TABLE 1 Guidelines for the Use of Uncertainty Factors in De riving Reference Doses and Modifying Factors

Standard Uncertainty Factors (UFs): Use a 10-fold factor when extrapolating from valid experimental results in studies using prolonged exposure to average healthy humans. This factor is intended to account for the variation in sensitivity among the members of the human population and is referenced as "10H". Use an additional 10-fold factor when extrapolating from valid results of longterm studies on experimental animals when results o f studies of human exposure are not available or are inadequate. This factor is intended to account for the uncertainty involved in extrapolating from animal d ata to humans and is referenced as "10A". Use an additional 10-fold factor when extrapolating from less than chronic results on experimental animals when there are no useful lo ng-term human data. This factor is intended to account for the uncertainty i nvolved in extrapolat- ing from less than chronic NOAELs to chronic NOAELs and is r eferenced as "10S". Use an additional 10-fold factor when deriving an R fD from a LOAEL, instead of a NOAEL. This factor is intended to account for the u ncertainty involved in extrapolating from LOAELs to NOAELs and is referenc ed as "10L". Modifying Factor (MF): 

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12/5/2011 http://www.epa.gov/iris/rfd.htmUse professional judgment to determine the MF, whic h is an additional uncertainty factor that is greater than zero and le ss than or equal to 10. The magnitude of the MF depends upon the professional a ssessment of scientific uncertainties of the study and data base not explic itly treated above; e.g., the completeness of the overall data base and the numbe r of species tested. The default value for the MF is 1.

*Source: Adapted from Dourson and Stara, 1983 Such practices, as employed by a number of scientis ts in different programs/agencies, exercising their best scientific judgment, have in some cases resulted in different ADIs for the same chemical. T he fact that different ADIs were generated (for example, by adopting different SFs) can be a source of considerable confusion when the ADIs are used exclu sively in risk management decisionmaking (see Section 1.2.2.2.3). The existen ce of different ADIs need not imply that any of them is more "wrong"--or "right"- -than the rest. It is more nearly a reflection of the honest difference in sci entific judgment. However, on occasion, these differences in judgment of the scientific data, can be interpreted as differences in the management of the risk. As a result, scientists may be inappropriately impugned, and/or perfectly j ustifiable risk management decisions may be tainted by charges of "tampering w ith the science." This unfortunate state of affairs arises, at least in pa rt, from treating the ADI as an absolute measure of safety. 1.3. EPA ASSESSMENT OF RISKS ASSOCIATED WITH SYSTEM IC TOXICITY The U.S. EPA approach to analyzing systemic toxicit y data follow the general format set forth by NRC in its description of the r isk assessment process (NRC, 1983). The determination of the presence of risk an d its potential magnitude is made during the risk assessment process, which cons ists of hazard identification, dose-response assessment, exposure assessment, and risk characterization. Having been apprised by the risk assessor that a po tential risk exists, the risk manager considers control options available under e xisting statutes and other relevant non-risk factors (e.g., benefits to be gai ned and costs to be incurred). All of these considerations go into the determination o f the regulatory decision (Figure 1).

FIGURE 1 Conceptual Framework for Risk Assessment and Risk M anagement*

--------------------------------------------------- --------------------------- Risk Assessment Risk Management

------------------------------------ -------- --------------------------- Hazard Identification Control Options | | 

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12/5/2011 http://www.epa.gov/iris/rfd.htm| |

| Dose-Response Assessment | Non -risk Analyses | (e.g., RfD) | | | | | | | | Exposure Assessment | | | | | | | | | | | |

v v v v v Risk Characterization ------------------->Regulato ry Decision (e.g., criterion) (e.g., RgD, Standard)

--------------------------------------------------- --------------------------- *Source: Adapted from NRC, 1983

1.3.1. HAZARD IDENTIFICATION 1.3.1.1. Evidence 1.3.1.1.1. Type of effect Exposure to a given chemical, depending on the dose employed, may result in a variety of toxic effects. These may range from gros s effects, such as death, to more subtle biochemical, physiologic, or pathologic changes. In assessments of the risk posed by a chemical, the toxic endpoints f rom all available studies are considered, although primary attention usually is g iven to the effect (the "critical effect") exhibiting the lowest NOAEL. In the case o f chemicals with limited data bases, additional toxicity testing may be necessary before an assessment can be made. 1.3.1.1.2. Principal studies Principal studies are those that contribute most si gnificantly to the qualitative assessment of whether or not a particular chemical is potentially a systemic toxicant in humans. In addition, they may be used i n the quantitative doseresponse assessment phase of the risk assessment. T hese studies are of two types: studies of human populations (epidemiologic investigations) and studies using laboratory animals. 1.3.1.1.2.1. Epidemiologic studies Human data are often useful in qualitatively establ ishing the presence of an adverse effect in exposed human populations. When t here is information on the exposure level associated with an appropriate endpo int, epidemiologic studies can also provide the basis for a quantitative dose- response assessment. The presence of such data obviates the necessity of ext rapolating from animals to 

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12/5/2011 http://www.epa.gov/iris/rfd.htmhumans; therefore, human studies, when available, a re given first priority, with

animal toxicity studies serving to complement them. In epidemiologic studies, confounding factors that are recognized can be controlled and measured, within limits. Case report s and acute exposures resulting in severe effects provide support for the choice of critical toxic effect, but they are often of limited utility in establishi ng a quantitative relationship between environmental exposures and anticipated eff ects. Available human studies on ingestion are usually of this nature. Co hort studies and clinical studies may contain exposure-response information that can be used in estimating effect levels, but the method of establishing exposure mus t be evaluated for validity and applicability. 1.3.1.1.2.2. Animal studies For most chemicals, there is a lack of appropriate information on effects in humans. In such cases, the principal studies are dr awn from experiments conducted on nonhuman mammals, most often the rat, mouse, rabbit, guinea pig, hamster, dog, or monkey. 1.3.1.1.3. Supporting studies These studies provide supportive, rather than defin itive, information and can include data from a wide variety of sources. For ex ample, metabolic and other pharmacokinetic studies can provide insights into t he mechanism of action of a particular compound. By comparing the metabolism of the chemical exhibiting the toxic effect in the animal with the metabolism foun d in humans, it may be possible to assess the potential for toxicity in hu mans or to estimate the equitoxic dose in humans. Similarly, in vitro studies can provide insights in to the chemical's potential for biological activity; and under certain circumstance s, considera- tion of structureactivity relationships between a chemical and other structurally related compounds can provide clues to the test chemical's pos- sible toxicity. More reliable in vitro tests are presently being develop ed to minimize the need for liveanimal testing. There is also increased emphasis on generating mechanism-ofaction and pharmacokinetic information as a means o f increasing understanding of toxic processes in humans and nonhumans. 1.3.1.1.4. Route of exposure The U.S. EPA often approaches the investigation of a chemical with a particular route of exposure in mind (e.g., an oral exposure f or a drinking water contaminant or an inhalation exposure for an air co ntaminant). In most cases, the toxicologic data base does not include detailed testing on all possible routes of administration, with their possibly significant differences in factors such as mechanism-of-action and bioavailability. In general , the U.S. EPA's position is that the potential for toxicity manifested via one route of exposure is relevant to considerations of any other route of exposure, unle ss convincing evidence exists to the contrary. Consideration is given to potentia l differences in absorption or metabolism resulting from different routes of expos ure, and whenever 

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12/5/2011 http://www.epa.gov/iris/rfd.htmappropriate data (e.g., comparative metabolism stud ies) are available, the

quantitative impacts of these differences on the ri sk assessment are delineated. 1.3.1.1.5. Length of exposure The U.S. EPA is concerned about the potential toxic effects in humans associated with all possible exposures to chemicals. The magni tude, frequency, and duration of exposure may vary considerably in different situ ations. Animal studies are conducted using a variety of exposure durations (e. g., acute, subchronic, and chronic) and schedules (e.g., single, intermittent, or continuous dosing). Information from all these studies is useful in the hazard identification phase of risk assessment. For example, overt neurological pr oblems identified in high-dose acute studies tend to reinforce the observation of subtle neurological changes seen in low-dose chronic studies. Special attention is given to studies involving low-dose, chronic exposures, since such exposures c an elicit effects absent in higher dose, shorter exposures, through mechanisms such as accumulation of toxicants in the organisms. 1.3.1.1.6. Quality of the study Evaluation of individual studies in humans and anim als requires the consideration of several factors associated with a study's hypoth esis, design, execution, and interpretation. An ideal study addresses a clearly delineated hypothesis, follows a carefully prescribed protocol, and includes suffici ent subsequent analysis to support its conclusions convincingly. In evaluating the results from such studies, consid eration is given to many other factors, including chemical characterization of the compound(s) under study, the type of test species, similarities and differences between the test species and humans (e.g., chemical absorption and metabolism), the number of individuals in the study groups, the number of study groups, the s pacing and choice of dose levels tested, the types of observations and method s of analysis, the nature of pathologic changes, the alteration in metabolic res ponses, the sex and age of test animals, and the route and duration of exposure. 1.3.1.2. Weight-of-Evidence Determination As the culmination of the hazard identification ste p, a discussion of the weight-ofevidence summarizes the highlights of the informati on gleaned from the principal and supportive studies. Emphasis is given to examin ing the results from different studies to determine the extent to which a consiste nt, plausible picture of toxicity emerges. For example, the following factors add to the weight of the evidence that the chemical poses a hazard to humans: similar results in replicated animal studies by different investigators; similar effects across sex, strain, species, and route of exposure; clear evidence of a dose-respons e relationship; a plausible relation between data on metabolism, postulated mec hanism-of-action, and the effect of concern; similar toxicity exhibited by st ructurally related compounds; and some link between the chemical and evidence of the effect of concern in humans. 1.3.2. DOSE-RESPONSE ASSESSMENT 

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12/5/2011 http://www.epa.gov/iris/rfd.htm1.3.2.1. Concepts and Problems Empirical observations have generally revealed that as the dosage of a toxicant is increased, the toxic response (in terms of severity and/or incidence of effect) also increases. This dose-response relationship is well- founded in the theory and practice of toxicology and pharmacology. Such behav ior is observed in the following instances: in quantal responses, in which the proportion of responding individuals in a population increases with dose; in graded responses, in which the severity of the toxic response within an individual increases with dose; and in continuous responses, in which changes in a biologi cal parameter (e.g., body or organ weight) vary with dose. In evaluating a dose-response relationship, certain difficulties arise. For example, one must decide on the critical endpoint to measure as the "response." One must also decide on the correct measure of "dose." In ad dition to the interspecies extrapolation aspects of the question of the approp riate units for dose, the more fundamental question of administered dose versus ab sorbed dose versus target organ dose should be considered. These questions ar e the subject of much current research. 1.3.2.2. Selection of the Critical Data 1.3.2.2.1. Critical study Data from experimental studies in laboratory animal s are often selected as the governing information when performing quantitative risk assessments, since available human data are usually insufficient for t his purpose. These animal studies typically reflect situations in which expos ure to the toxicant has been carefully controlled and the problems of heterogene ity of the exposed population and concurrent exposures to other toxicants have be en minimized. In evaluating animal data, a series of professional judgments are made which involve, among others, consideration of the scientific quality of the studies. Presented with data from several animal studies, the risk assessor firs t seeks to identify the animal model that is most relevant to humans, based on the most defensible biological rationale (e.g., for instance using comparative pha rmacokinetic data). In the absence of a clearly most relevant species, the mos t sensitive species (i.e., the species showing a toxic effect at the lowest admini stered dose) is used by risk assessors at U.S. EPA, since there is no assurance that humans are not at least as innately sensitive as the most sensitive species tested. This selection process is more difficult when the routes of exposure in th e animal tests are different from those involved in the human situation under in vestigation. In order to use data from controlled studies of genetically homogen eous animals, the risk assessor must also extrapolate from animals to huma ns and from high experimental doses to comparatively low environment al exposures, and must account for human heterogeneity and possible concur rent human exposures to other chemicals. Although for most chemicals there is a lack of well -controlled cohort studies investigating noncancer endpoints, in some cases an epidemiologic study may be selected as the critical data (e.g., in cases of ch olinesterase inhibition). Risk assessments based on human data have the advantage of avoiding the problems inherent in interspecies extrapolation. In many ins tances, use of such studies, as 

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12/5/2011 http://www.epa.gov/iris/rfd.htmis the case with animal investigations, involves ex trapolation from relatively high

doses (such as those found in occupational settings ) to the low doses found in the environmental situations to which the general p opulation is more likely to be exposed. In some cases, a well-designed and well-co nducted epidemiologic study that shows no association between known exposures a nd toxicity can be used to directly project an RfD (as has been done in the ca se of fluoride). 1.3.2.2.2. Critical data In the simplest terms, an experimental exposure lev el is selected from the critical study that represents the highest level tested in w hich "no adverse effect" was demonstrated. This "no-observed-adverse-effect-leve l" (NOAEL) is the key datum gleaned from the study of the dose-response relatio nship and, traditionally, is the primary basis for the scientific evaluation of the risk posed to humans by systemic toxicants. This approach is based on the a ssumption that if the critical toxic effect is prevented, then all toxic effects a re prevented. More formally, the NOAEL is defined in this discuss ion as the highest experimental dose of a chemical at which there is n o statistically or biologically significant increase in frequency or severity of an adverse effect in individuals in an exposed group when compared with individuals in an appropriate control group. As noted above, there may be sound professio nal differences of opinion in judging whether or not a particular response is adv erse. In addition, the NOAEL is a function of the size of the population under stud y. Studies with a small number of subjects are less likely to detect low-dose effe cts than studies using larger numbers of subjects. Also, if the interval between doses in an experiment is large, it is possible that the experimentally deter mined NOAEL is lower than that which would be observed in a study using intervenin g doses. 1.3.2.2.3. Critical endpoint As noted in Section 1.2., a chemical may elicit mor e than one toxic effect (endpoint), even in one test animal, or in tests of the same or different duration (acute, subchronic, and chronic exposure studies). In general, NOAELs for these effects will differ. The critical endpoint used in the dose-response assessment is the effect exhibiting the lowest NOAEL. 1.3.2.3. Reference Dose (RfD) The reference dose (RfD) and uncertainty factor (UF ) concepts have been developed by the RfD Work Group in response to many of the problems associated with ADIs and SFs, as previously outline d in Section 1.2. The RfD is a benchmark dose operationally derived from the NOAEL by consistent application of generally order-of-magnitude uncertainty factors (UFs) that reflect various types of data sets used to estimate RfDs. For examp le, a valid chronic animal NOAEL is normally divided by an UF of 100. In addit ion, a modifying factor (MF), is sometimes used which is based on a professional judgment of the entire data base of the chemical. These factors and their ratio nales are presented in Table 1. The RfD is determined by use of the following equat ion: RfD = NOAEL / (UF x MF) 

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12/5/2011 http://www.epa.gov/iris/rfd.htmwhich is the functional equivalent of Equation 1. I n general, the RfD is an

estimate (with uncertainty spanning perhaps an orde r of magnitude) of a daily exposure to the human population (including sensiti ve subgroups) that is likely to be without an appreciable risk of deleterious effec ts during a lifetime. The RfD is generally expressed in units of milligrams per kilo gram of bodyweight per day (mg/kg/day). The RfD is useful as a reference point from which t o gauge the potential effects of the chemical at other doses. Usually, doses less th an the RfD are not likely to be associated with adverse health risks, and are there fore less likely to be of regulatory concern. As the frequency and/or magnitu de of the exposures exceeding the RfD increase, the probability of adve rse effects in a human population increases. However, it should not be cat egorically concluded that all doses below the RfD are "acceptable" (or will be ri sk-free) and that all doses in excess of the RfD are "unacceptable" (or will resul t in adverse effects). The U.S. EPA is attempting to standardize its appro ach to determining RfDs. The RfD Work Group has developed a systematic approach to summarizing its evaluations, conclusions, and reservations regardin g RfDs in a "cover sheet" of a few pages in length. The cover sheet includes a sta tement on the confidence (high, medium, or low) the evaluators have in the s tability of the RfD. High confidence indicates the judgment that the RfD is u nlikely to change in the future because there is consistency among the toxic respon ses observed in different sexes, species, study designs, or in dose-response relationships, or that the reasons for existing differences are well understoo d. High confidence is often given to RfDs that are based on human data for the exposure route of concern, since in such cases the problems of interspecies ex trapolation have been avoided. Low confidence indicates the judgment that the data supporting the RfD may be of limited quality and/or quantity and that additio nal information could result in a change in the RfD. 1.3.3. EXPOSURE ASSESSMENT The third step in the risk assessment process focus es on exposure issues. For a full discussion of exposure assessment, consult U.S . EPA's guidelines on the subject (U.S. EPA, 1987). In brief, the exposure as sessment includes consideration of the size and nature of the populat ions exposed and the magnitude, frequency, duration and routes of exposu re, as well as evaluation of the nature of the exposed populations. 1.3.4. RISK CHARACTERIZATION Risk characterization is the final step in the risk assessment process and the first input to the risk management (regulatory action) pr ocess. The purpose of risk characterization is to present the risk manager wit h a synopsis and synthesis of all the data that should contribute to a conclusion with regard to the nature and extent of the risk, including: a. The qualitative ("weight-of-evidence") conclusio ns as to the likelihood that the chemical may pose a hazard to human health. 

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12/5/2011 http://www.epa.gov/iris/rfd.htmb. A discussion of the dose-response information co nsidered in deriving the the RfD, including the UFs and MFs used. c. Data on the shapes and slopes of the dose-respon se curves for the various toxic endpoints, toxicodynamics (absorption and met abolism), structure-activity correlations, and the nature and severity of the ob served effects. d. Estimates of the nature and extent of the exposu re and the number and types of people exposed. e. Discussion of the overall uncertainty in the ana lysis, including the major assumptions made, scientific judgments employed, an d an estimate of the degree of conservatism involved. In the risk characterization process, a comparison is made between the RfD and the estimated (calculated or measured) exposure dos e (EED). The EED should include all sources and routes of exposure involved . If the EED is less than the RfD, the need for regulatory concern is likely to b e small. An alternative measure that may be useful to some r isk managers is the margin of exposure (MOE), which is the magnitude by which the NOAEL of the critical toxic effect exceeds the estimated exposure dose (E ED), where both are expressed in the same units: MOE = NOAEL (experimental dose) / EED (human dose). When the MOE is equal to or greater than UF x MF, t he need for regulatory concern is likely to be small. Section 1.6. contains an example of the use of the concepts of NOAEL, UF, MF, RfD, EED, and MOE. 1.4. APPLICATION IN RISK MANAGEMENT Once the risk characterization is completed, the fo cus turns to risk management. In reaching decisions, the risk manager utilizes th e results of risk assessment, other technological factors, and legal, economic an d social considerations in reaching a regulatory decision. These additional fa ctors include efficiency, timeliness, equity, administrative simplicity, cons istency, public acceptability, technological feasibility, and nature of the legisl ative mandate. Because of the way these risk management factors ma y impact different cases, consistent -- but not necessarily identical -- risk management decisions must be made on a case-by-case basis. For example, the Clea n Water Act calls for decisions with "an ample margin of safety"; the Fed eral Insecticide, Fungicide and Rodenticide Act (FIFRA) calls for "an ample mar gin of safety," taking benefits into account; and the Safe Drinking Water Act (SDWA ) calls for standards which protect the public "to the extent feasible." Conseq uently, it is entirely possible and appropriate that a chemical with a specific RfD may be regulated under different statutes and situations through the use o f different "regulatory doses (RgDs)." 

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12/5/2011 http://www.epa.gov/iris/rfd.htmThat is, after carefully considering the various ri sk and nonrisk factors, regulatory

options, and statutory mandates in a given case (i) , the risk manager selects the appropriate statutory alternative for arriving at a n "ample" or "adequate" margin of exposure [MOE(i)]. As shown in Equation 2 below, this procedure establishes the regulatory dose, RgD(i) (e.g., a tolerance unde r FIFRA or a maximum contaminant level under SDWA), applicable to the ca se in question: RgD(i) = NOAEL / MOE(i). (Equation 2) Note that different RgDs are possible for a given c hemical with a single RfD. Note also that comparing the RfD to a particular RgD(i) is equivalent to comparing the MOE(i) with the UF x MF: RfD/RgD(i) = MOE(i) / (UF x MF). In assessing the significance of a case in which th e RgD is greater (or less) than the RfD, the risk manager should carefully consider the case-specific data compiled by the risk assessors, as discussed in Sec tion 1.3.4. In some cases, additional explanation and interpretation may be re quired from the risk assessors in order to arrive at a responsible and clearly art iculated final decision on the RgD. It is generally useful to the risk manager to have information regarding the contribution to the RfD from various environmental media (e.g., air, water and food). Such information can provide insights that a re helpful in choosing among available control options. However, in cases in whi ch site-specific criteria are being considered, local exposures through various m edia can often be determined more accurately than exposure estimates based upon generic approaches. In such cases, the exposure assessor's role is particu larly important. For instance, at a given site, consumption of fish may clearly domin ate the local exposure routes, while, on a national basis, fish consumption may pl ay a minor role compared to ingestion of treated crops. Work is underway in the U.S. EPA to apportion the R fD among the various environmental media. For example, consider the case of a food-use pesticide which is a contaminant in drinking water. In select ing among risk management actions under the Safe Drinking Water Act, it might be prudent to assume an RfD for drinking water purposes which is some fraction of the total RfD. Such an apportionment would explicitly acknowledge the poss ible additional exposure from ingestion of treated crops. The apportionment of the RfD would, in general, provide additional guidance for risk managers of th e various media-specific programs. 1.5. OTHER DIRECTIONS In addition to the development of reference doses, the U.S. EPA is pursuing other lines of investigation for systemic toxicity. For e xample, the Office of Air Quality Planning and Standards is using probabilistic risk assessment procedures for criteria pollutants. In this procedure, the populat ion at risk is characterized, and the likelihood of the occurrence of various effects is predicted through the use of available scientific literature and of scientific e xperts' rendering their judgments concerning dose-response relationships. This dose-r esponse information is then 

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12/5/2011 http://www.epa.gov/iris/rfd.htmcombined with the results of the exposure analysis to generate population risk

estimates for alternative standards. Through the us e of these procedures, decisionmakers are presented with ranges of risk es timates in which uncertainties associated with both the toxicity and exposure info rmation are explicitly considered. The Office of Policy, Planning and Eval uation is investigating similar procedures in order to balance health risk and cost . In addition, scientists in the Office of Research and Development have initiated a series of studies designed to increase the reliability of risk assessments. They are investigating the use of extrapolation models as a means of estimating RfDs, taking into account the statistical variability of the NOAEL and underlying UFs. ORD is also exploring procedures for conducting health risk assessments t hat involve less-than-lifetime exposures. Finally, they are working on approaches to ranking the severity of different toxic effects. 1.6. HYPOTHETICAL, SIMPLIFIED EXAMPLE OF DETERMININ G AND USING RfD 1.6.1. EXPERIMENTAL RESULTS Suppose the U.S. EPA had a sound 90-day subchronic gavage study in rats with the data in Table 2:
TABLE 2

Hypothetical Data to Illustrate the Reference Dose Concept

--------------------------------------------------- --------------------------- Dose Observation Effect Level mg/kg/day

--------------------------------------------------- --------------------------- 0 Control--no adverse effects observe d --

1 No statistically or biologically NOEL significant differences between tre ated
and control animals

5 2% decrease* in body weight gain (n ot NOAEL considered to be of biological sign ificance); increased ratio of liver weight to body weight; histopathology indistinguishable fr om controls; elevated liver enzyme levels

25 20% decrease* in body weight gain; LOAEL 

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12/5/2011 http://www.epa.gov/iris/rfd.htmincreased* ratio of liver weight to body

weight; enlarged, fatty liver with vacuole formation; increased* liver enzyme levels

--------------------------------------------------- --------------------------- *Statistically significant compared to controls. 1.6.2. ANALYSIS 1.6.2.1. Determination of the Reference Dose (RfD) 1.6.2.2.1. Using the NOAEL Because the study is on animals and of subchronic d uration, UF = 10H x 10A x 10S = 1000 (Table 1). In addition, there is a subjective adjustment (MF) based on the high number of animals (250) per dose group: MF = 0.8. These factors then give UF x MF = 800, so that RfD = NOAEL/(UF x MF) = 5/800 = 0.006 (mg/kg/day). 1.6.2.1.2. Using the LOAEL If the NOAEL is not available, and if 25 mg/kg/day had been the lowest dose tested that showed adverse effects, UF = 10H x 10A x 10S x 10L = 10,000 (Table 1). Using again the subjective adjustment of MF = 0.8, one obtains RfD = LOAEL/(UF x MF) = 25/8000 = 0.003 (mg/kg/day)

. 1.6.2.2. Risk Characterization Considerations Suppose the estimated exposure dose (EED) for human s exposed to the chemical under the proposed use pattern were 0.01 mg/kg/day (i.e., the EED is greater than the RfD). Viewed alternatively, the MOE is: MOE = NOAEL/EED = 5 (mg/kg/day) / 0.01 (mg/kg/day) = 500. Because the EED exceeds the RfD (and the MOE is les s than the UF x MF), the risk manager will need to look carefully at the dat a set, the assumptions for both the RfD and the exposure estimates, and the comment s of the risk assessors. In 

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12/5/2011 http://www.epa.gov/iris/rfd.htmaddition, the risk manager will need to weigh the b enefits associated with the

case, and other non-risk factors, in reaching a dec ision on the regulatory dose (RgD). 1.7. REFERENCES Dourson, M.L. and J.F. Stara. 1983. Regulatory Toxi cology and Pharmacology. 3: 224-238. Lehman, A.J. and O.G. Fitzhugh. 1954. Association o f Food Drug Officials. USQ Bull. 18: 33-35. NRC (National Research Council). 1983. Risk Assessm ent in the Federal Government: Managing the Process. NAS Press, Washin gton, DC. U.S. EPA. 1987. The Risk Assessment Guidelines of 1 986. Office of Health and Environmental Assessment, Washington, DC. EPA/600/8 -87/045. http://www.epa.gov/iris/rfd.htm Last updated on Tuesday, July 26, 2011 

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12/5/2011 http://www.epa.gov/iris/rfd.htm