Risk Methods



Cancer Risk Assessment Methods Classification Schemes


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Dose-Response Assessment
 
bullet ARA Dose-Response Framework
bullet ATSDR
bullet Health Canada
bullet IARC
bullet TCEQ
bullet NSF International
bullet RIVM
bullet U.S. EPA

References

 

 

 

Noncancer Risk Assessment Methods

 

bullet ATSDR
bullet Health Canada
bullet NSF International
bullet RIVM
bullet TCEQ
bullet U.S. EPA

 

References


Cancer Classification Schemes

ATSDR

 

ATSDR's qualitative conclusions regarding carcinogenicity are presented in the Toxicological Profiles using a weight of evidence approach. This approach relies upon NTP's Annual Report on Carcinogens. Conclusions from IARC, U.S. EPA and OSHA are also presented as appropriate.

 


 

Health Canada

 

Health Canada classifies chemicals into six categories with regard to carcinogenicity based on a modification of the scheme used by the International Agency for Research on Cancer. The following is excerpted from Human Health Risk Assessment for Priority Substances (Health Canada, 1994):

 

Group I: Carcinogenic to Humans

Data from adequate epidemiological studies indicate that there is a causal relationship between exposure to a substance and an increased incidence of cancer in humans.

 

Group II: Probably Carcinogenic to Humans

Data from epidemiological studies are inadequate to assess carcinogenicity. However, there is sufficient evidence of carcinogenicity in animal species (i.e., there is an increased incidence of malignant tumours in multiple species or strains, in multiple experiments with different routes of exposure or dose levels, or the incidence, site or type of tumour or age of onset is unusual). Exceptionally, a compound for which the evidence of carcinogenicity is limited but for which there is a strong supporting dataset (on genotoxicity, for example) which indicates that the compound is likely to be carcinogenic would be included in this category.

 

Group III: Possibly Carcinogenic to Humans

Health Canada has three subgroups in Group III which describe the data in humans and laboratory animals that would result in a classification in this Group. In summary, this includes chemicals for which data from epidemiological studies are inadequate, or which indicate an association between exposure and human cancer but alternative explanations such as chance, bias or confounding cannot be excluded. There is some evidence of increased tumour incidence in animals but the data are limited because the studies involve a single species, strain or experiment; study design (i.e., dose levels, duration of exposure and follow-up, survival, number of animals) or reporting is inadequate; the neoplasms produced often occur spontaneously and have been difficult to classify as malignant by histological criteria alone (e.g., lung and liver tumours in mice). The weight of limited evidence indicates that the compound is genotoxic or results are mixed. Chemicals believed to have an epigenetic mechanism of cancer induction may also be classified in Group III if there are positive cancer studies in long-term animal experiments.

 

Group IV: Unlikely to be Carcinogenic to Humans

Health Canada has four subgroups in Group IV which describe the data in humans and laboratory animals that would result in a classification in this Group. In summary, this includes chemicals for which there is no evidence of carcinogenicity in adequate epidemiological studies or data are inadequate. There is some evidence of carcinogenicity in well -designed and well-conducted carcinogenicity bioassays in animals, but the results are limited or can be confidently ascribed to species-specific mechanisms of toxicity and/or metabolism which do not appear to be operative in humans.

 

Group V: Probably Not Carcinogenic to Humans

Health Canada has three subgroups in Group V which describe the data in humans and laboratory animals that would result in a classification in this Group. In summary, this includes chemicals for which there is no evidence of carcinogenicity in sufficiently powerful and well-designed epidemiological studies; there is no evidence or inadequate data on carcinogenicity in laboratory animals.

 

Group VI: Unclassifiable with Respect to Carcinogenicity in Humans

Health Canada has three subgroups in Group VI which describe the data in humans and laboratory animals that would result in a classification in this Group. In summary, this includes chemicals for which data from epidemiological and/or animal studies are inadequate or not available.

 


 

IARC

An explanation of IARC's methods is available in the Preamble to the IARC Monographs at http://monographs.iarc.fr/ENG/Preamble/index.php. The Preamble to the Monographs sets out the objective and scope of the evaluation programme, the procedures used when making assessments, and the types of evidence considered and criteria used in reaching the final evaluations.

 


 

NSF International

NSF International currently uses the U.S. EPA (2005) weight of evidence narrative approach to cancer classification. The conclusion reached by NSF is included as part of the hazard assessment in a weight of evidence evaluation and cancer characterization section of the oral risk assessment document. If the U.S. EPA or another internationally recognized organization such as the NTP (National Toxicology Program), ATSDR, Health Canada, IARC or other members of the World Health Organization has also classified the chemical, that classification will be included in the risk comparisons and conclusions section of the NSF document, with discussion if the classifications differ. The U.S. EPA and NSF International classifications may occasionally differ if new data have been evaluated by one of the organizations.

 


 

RIVM

An explanation of RIVM's risk assessment methods is available in the following report:

Janssen, PJCM and GJA Speijers. 1997. Guidance on the Derivation of Maximum Permissible Risk Levels for Human Intake of Soil Contaminants. Report no. 711701006, National Institute of Public Health and the Environment. Bilthoven, The Netherlands. January. Available at http://www.rivm.nl/bibliotheek/rapporten/711701006.pdf or at http://www.rivm.nl/en/ (click on Search, type "711701006", then click on document).

 


 

TCEQ

An explanation of TCEQ's methods is available in the publication entitled, TCEQ Guidelines to Develop Toxicity Factors” Available at http://www.tceq.texas.gov/publications/rg/rg-442.html .  This document is a technical guide that details the process of developing Effects Screening Levels (ESLs), Reference Values (ReVs), and Unit Risk Factors (URFs).

 

 


 

 

U.S. EPA

In 1986, the U.S. EPA published general guidelines to be used by Agency scientists in developing and evaluating risk assessments for carcinogens (U.S. EPA, 1986). Almost all of the carcinogen assessments on IRIS were based on the 1986 guidelines. Assessments developed between 1996 and approximately 1999 may have used the 1996 proposed guidelines; and assessments developed between approximately 1999 and early 2005 may have used the 1999 draft guidelines. Both the 1996 and 1999 versions were similar to the 2005 (final) version, and used comparable quantitative approaches. However, the 1996 version included two fewer categories, and both 1996 and 1999 versions differed in some other details from the 2005 guidelines. For more details about the evolution of U.S. EPA’s cancer guidelines, please see http://www.epa.gov/cancerguidelines/.


Below is a brief description of the weight of evidence and cancer classification guidelines from the 1986 guidelines. This is followed by a brief description of the 2005 guidelines.

 

Description of 1986 Guidelines

The 1986 guidelines specify that information be categorized into one of three types: human data, animal data, and supporting data. The human and animal data are used to make a preliminary judgment as to the likelihood that the agent in question may produce tumors in humans. The supporting data (e.g., genotoxicity, mechanistic data, and pharmacokinetic information) are then used to elevate or downgrade the classification. For a description of the amount and type of data required for a chemical to be assigned to any one of these groups, the reader is referred to the 1986 guidelines (U.S. EPA, 1986). In brief, the categories from the 1986 guidelines, as defined by U.S. EPA, are as follows:

 

Group A: Carcinogenic to humans

Classification in Group A requires the observation of a statistically significant association between exposure to an agent and malignant or life-threatening benign tumors in humans.

 

Group B: Probably carcinogenic to humans

EPA divides this group into the categories B1 and B2. Limited human evidence of carcinogenicity in humans is necessary for placement of a chemical in Group B1. Group B2 includes chemicals with sufficient animal evidence, but inadequate human evidence for carcinogenicity.

 

Group C: Possibly carcinogenic to humans

An agent is classified in Group C when human data are inadequate and animal data demonstrate limited evidence of carcinogenicity (e.g., an increased incidence of benign tumors only; a positive finding of carcinogenicity in one species only; an increased incidence of neoplasms that occur with high spontaneous background incidence)

 

Group D: Not classifiable as to human carcinogenicity

An agent is classified in Group D when insufficient data are available to make a determination as to carcinogenicity.

 

Group E: Evidence of noncarcinogenicity for humans

An agent is classified in Group E if there is no increased incidence of neoplasms in at least two well-designed and well-conducted animal studies of adequate power and dose in different species.

 

Description of 2005 Guidelines

The 2005 guidelines (U.S. EPA, 2005) significantly change the way hazard evidence is weighed in reaching conclusions about an agent's potential for human carcinogenicity. Tumor findings in animals or humans dominated the 1986 classification scheme. Under the 2005 guidelines, decisions are based on all of the evidence, particularly information regarding mode of action at cellular and subcellular levels, as well as toxicokinetics and metabolic processes. Weighing of the evidence includes considering the likelihood of human carcinogenic effects of the agent and the conditions under which such effects may be expressed, as these are revealed in the toxicological and other biologically important features of the agent. This more complete characterization of the expression of carcinogenic potential might include findings that an agent is observed to be carcinogenic by one route, but not another. Alternatively, the agent's carcinogenic activity might be secondary to another toxic effect.

 

The 2005 guidelines use standard descriptors of conclusions rather than letter designations. The descriptors are incorporated into a brief narrative that explains an agent’s human carcinogenic potential and the conditions that characterize its expression. Significant issues, strengths, and limitations of the data and conclusions are included. The narrative also summarizes the mode of action information that underlies the approach to dose-response assessment. Five categories of descriptors are used, with additional text further defining the conclusion. In brief, the descriptors from the 2005 guidelines are:

 

“Carcinogenic to Humans”

This descriptor is appropriate when there is convincing epidemiologic evidence demonstrating causality between human exposure and cancer. EPA also considers this descriptor to be appropriate when there is an absence of conclusive epidemiologic evidence to clearly establish a cause and effect relationship between human exposure and cancer, but a number of other criteria are met. The criteria are (1) strong evidence of an association between human exposure and either cancer or key precursor events, (2) extensive evidence of carcinogenicity in animals, (3) the mode(s) of action and key precursor events have been identified in animals, and (4) there is strong evidence that the key precursor events are anticipated to occur in humans and progress to tumors.

 

“Likely to be Carcinogenic to Humans”

This descriptor is appropriate when the available tumor effects and other key data are adequate to demonstrate carcinogenic potential to humans. Adequate data are within a spectrum. At one end is evidence for a plausible (but not definitively causal) association between human exposure to the agent and cancer, usually with some supporting evidence (not necessarily carcinogenicity data) in animals. At the other end of the spectrum is an agent with no human data, but a positive tumor study in animals and the weight of experimental evidence shows that in experimental animals the agent causes events generally known to be associated with tumor formation.

 

“Suggestive Evidence of Carcinogenic Potential”

This descriptor is appropriate when the weight of evidence from human or animal data is suggestive of carcinogenicity; a concern for carcinogenic effects in humans is raised, but is judged not sufficient for a stronger conclusion. Examples of such evidence may include: (1) a small and possibly not statistically significant increase in tumors in a single study that is not contradicted by other studies of equal quality in the same system, or (2) a small increase in a tumor with a high background rate in that sex and strain, when there is some evidence that the observed tumors may be due to intrinsic factors. Dose-response assessment is generally not indicated for these agents.

 

“Data are Inadequate for an Assessment of Human Carcinogenic Potential”

This descriptor is used when available data are judged inadequate to perform an assessment. This includes a case when there is a lack of pertinent or useful data or when existing evidence is conflicting, e.g., some evidence is suggestive of carcinogenic effects, but other studies of equal quality in the same sex and strain are negative.

 

“Not Likely to be Carcinogenic to Humans”

This descriptor is used when the available data are considered robust for deciding that there is not basis for human hazard concern. This judgment may be based on (1) animal evidence that demonstrates lack of carcinogenic effect in at least two well-designed and well-conducted studies in two appropriate animal species (in the absence of other animal or human data suggesting a potential for cancer effects); (2) extensive experimental evidence showing that the only carcinogenic effects observed in animals are not considered relevant to humans; (3) convincing evidence that carcinogenic effects are not likely by a particular dose route; or (4) convincing evidence that carcinogenic effects are not anticipated below a defined dose range.

Cancer Dose Response Methods

ATSDR

ATSDR does not currently perform dose-response assessments for carcinogens. This Agency does, however, report values established by other Agencies (e.g., U.S. EPA, IARC).

ATSDR does not currently engage in low-dose modeling efforts or in the development of cancer potency factors (ATSDR 1993).

 


 

Health Canada

For substances considered by Health Canada to have no threshold (i.e., mutagens and genotoxic carcinogens), it is assumed that there is some probability of harm to human health at any level of exposure. For these chemicals, Health Canada considers it inappropriate to specify a concentration or dose associated with a negligible or de minimis level of risk (e.g., the1 in a million risk often used by U.S. EPA) by low-dose extrapolation procedures. Rather, potency is expressed as the dose or concentration which induces a 5% increase in the incidence of, or deaths due to, tumours or heritable mutations considered to be associated with exposure. The TD05/TC05 is then compared with exposure levels. If the ratio between exposure and the TD05/TC05 is less than 2 x 10-6, there is little need for further action. If the ratio is 2 x 10-4 or greater, there is a high priority for further action. Values in between are of moderate priority.

 

In order to compare cancer potencies estimated by different Agencies, TERA chose to express each Agency's potency value as the equivalent of a 1 in a 100,000 risk level. For Health Canada, this required dividing the TD05/TC05 (i.e., a 1 in 20 risk level) by a factor of 5,000 to represent a 1 in a 100,000 risk level. It is noted, however, that unlike the methodology used by U.S. EPA, Health Canada's TD05/TC05 is not based on a confidence limit, but is computed directly from the dose-response curve within or close to the experimental range. Health Canada considered this to be appropriate in view of the stability of the data in the experimental range and to avoid unnecessarily conservative assumptions.

 


 

IARC

An explanation of IARC's methods is available in the Preamble to the IARC Monographs at http://monographs.iarc.fr/ENG/Preamble/index.php. The Preamble to the Monographs sets out the objective and scope of the evaluation programme, the procedures used when making assessments, and the types of evidence considered and criteria used in reaching the final evaluations.

 


 

NSF International

NSF International currently uses U.S. EPA (2005) dose-response assessment methodology. Earlier documents have used U.S. EPA (1999) draft, U.S. EPA (1996) proposed or U.S. EPA (1986) final guidelines. Specific implementation of this methodology is described in Annex A of NSF International/American National Standard 60 “Drinking water treatment chemicals – Health effects,” and of NSF International/American National Standard 61 “Drinking water system components – Health effects.” For a tumor endpoint, human equivalent doses are first calculated by scaling the applied daily doses to body weight raised to the 0.25 power. The dose-response data are then subject to benchmark dose modeling (U.S. EPA, 2009) to determine the point of departure, which is generally the 95% confidence limit on a dose associated with an estimated 10% increased tumor or related non-tumor response (the LED10). If the weight of evidence suggests that the compound is genotoxic, the dose-response assessment is performed by linear extrapolation from the point of departure to a specific risk level. If there is a plausible mode of action that indicates the tumor or tumor precursor is not produced by a genotoxic mechanism, a margin-of-exposure approach may be used. In this latter case, if the data cannot be modeled, a NOAEL or LOAEL may be used as the point of departure.

 


 

RIVM

An explanation of RIVM's risk assessment methods is available in the following report:

 

Janssen, PJCM and GJA Speijers. 1997. Guidance on the Derivation of Maximum Permissible Risk Levels for Human Intake of Soil Contaminants. Report no. 711701006, National Institute of Public Health and the Environment. Bilthoven, The Netherlands. January. Available at http://www.rivm.nl/bibliotheek/rapporten/711701006.pdf or at http://www.rivm.nl/en/ (click on Search, type "711701006", then click on document).

 


 

TCEQ

An explanation of TCEQ's methods is available in the publication entitled, TCEQ Guidelines to Develop Toxicity Factors” Available at http://www.tceq.texas.gov/publications/rg/rg-442.html .  This document is a technical guide that details the process of developing Effects Screening Levels (ESLs), Reference Values (ReVs), and Unit Risk Factors (URFs).

 


 

U.S. EPA

U.S. EPA published guidelines for carcinogen risk assessment in 1986 (U.S. EPA, 1986). These guidelines outline procedures for estimating cancer potency. Almost all of the carcinogen assessments on IRIS were based on these 1986 guidelines. In 1996, EPA proposed revisions to the cancer guidelines (U.S. EPA, 1996), and these were further modified in the draft 1999 guidelines (U.S. EPA, 1999), and were then finalized in the 2005 guidelines (U.S. EPA, 2005). Assessments developed between 1996 and approximately 1999 may have used the 1996 proposed guidelines; and assessments developed between approximately 1999 and early 2005 may have used the 1999 draft guidelines. For more details about the evolution of U.S. EPA’s cancer guidelines, please see http://www.epa.gov/cancerguidelines/.

 

Below is a brief description of EPA dose-response procedures based on the 1986 guidelines and an explanation of how TERA expresses the results on ITER for comparison purposes. This is followed by a brief description of the 2005 guidelines.

 

Description of 1986 Guidelines

Two extrapolations are generally necessary when using animal data. The first step is extrapolation from animals to humans. According to the 1986 guidelines, this extrapolation is done by estimating a human equivalent oral dose, by scaling the daily applied doses to body weight raised to the 0.66 power. Second, one needs to extrapolate from the high doses used in animal studies to the generally lower doses of interest for environmental exposure. Risk at low exposure levels generally cannot be measured directly (either by animal experiments or by epidemiologic studies). Therefore, a number of mathematical models and procedures have been developed for use in extrapolating from high to low doses. Under EPA's 1986 cancer risk assessment guidelines, the linearized multistage model was generally chosen as the default model for extrapolation to low doses. Multistage models are exponential models approaching 100% risk at high doses, with a shape at low doses described by a polynomial function. The multistage model is fit to the tumor dose-response data, and an upper bound for the risk is estimated by incorporating an appropriate linear term into the statistical bound for the polynomial. At sufficiently small exposures, any higher-order terms in the polynomial will contribute negligibly, and the graph of the upper bound will appear to be a straight line. The slope of this line (formerly called the potency) is called the slope factor. Its units are (proportion of individuals with tumors)/mg/kg-day.

 

For the oral route, EPA calculates both a slope factor and a unit risk. As described above, the oral slope factor expresses the risk per mg/kg-day. The unit risk is a numerically equivalent term that is expressed as the risk associated with a drinking water concentration of 1 ug/L (with assumptions being made that an adult weighs 70 kg and drinks 2 L/day). For the route of inhalation, EPA does not provide a slope factor, but rather expresses the risk only in terms of a unit risk. The units for the inhalation unit risk are risk per 1 ug/m3. In other words, it is the risk associated with an air concentration of 1 ug/m3 (assuming a 70 kg adult breathes 20 cubic meters/day).

 

In order to compare cancer potencies estimated by different Agencies, TERA chose to express each Agency's potency value as the equivalent of a 1 in a 100,000 risk level. Thus, TERA calculates risk specific doses (RSDs) from EPA's oral slope factors and risk specific concentrations (RSCs) from EPA’s inhalation unit risks. Specifically, for oral slope factors, TERA converts the EPA risk estimate to a concentration at the 1 in 100,000 (E-5) risk level by dividing 1E-5 by the unit risk [in units of “per (ug/m3)”] and then by another 1000 to convert to mg/cu.m to determine a risk specific concentration (RSC) (in units of “mg/ m3”). Similarly, TERA converts the EPA oral slope factors to a dose at the 1 in 100,000 (E-5) risk level by dividing 1E-5 by the slope factor [in units of “per (mg/kg-day”] to determine a risk specific dose (RSD) (in units of “mg/kg-day”).

 

In setting standards for carcinogens, EPA generally considers a de minimis (e.g., less than or equal to 1 in a million) risk to be an acceptable goal. Using the output from the linearized multistage model, EPA often determines the oral intake or inhalation concentration that is associated with a risk of 1 in a million as a goal for setting limits on exposure. Risk management issues may lead to the setting of intakes/concentrations that are higher or lower.

 

Description of 2005 Guidelines

The 2005 cancer guidelines (U.S. EPA, 2005) differ significantly from the 1986 guidelines. When animal studies are used, the estimation of a human equivalent dose utilizes toxicokinetic models when available, and if not, the default for oral doses is to scale the daily applied doses to body weight raised to the 0.75 power. The default dose scaling methodology for inhalation follows that developed for derivation of reference concentrations (RfCs), estimating the relative animal and human respiratory deposition of particles, and the relative internal dose or dose to the respiratory region of gases, depending on the chemical and physical properties of the gas.

 

Response data from effects of the agent on carcinogenic processes (i.e., nontumor data) are analyzed along with tumor incidence data. Tumor incidences and precursor effects may be combined to extend the dose-response curve below the tumor data. A biologically based or case-specific dose response model to relate dose and response data in the range of empirical observation may be used when data are sufficient. When this is not the case, standard default procedures are used to fit a curve to the data and to calculate the lower 95% confidence limit on a dose associated with an estimated 10% increased tumor or relevant nontumor response (LED10). The LED10 then serves as a point of departure for extrapolating outside the observable range. Depending on the mode(s) of action of the agent, low-dose extrapolation from the LED10 is done using a linear approach, a nonlinear approach, or both. Linear extrapolation to low doses is used when the mode of action data indicates that the agent is DNA-reactive and has direct mutagenic activity, or if the human exposure or body burden is high and near doses associated with key precursor events. Linear extrapolation is also used as a default when there are insufficient data to evaluate mode of action. For linear extrapolation, a straight line is drawn from the point of departure to zero dose, zero response, corrected for background. The slope of the line expresses the extra risk per unit dose. This risk can be converted to the risk specific dose or risk specific concentration, as described for the 1986 guidelines. A nonlinear extrapolation is used when there a tumor mode of action supporting nonlinearity applies, and the chemical does not demonstrate mutagenic effects consistent with linearity. Alternatively, a nonlinear extrapolation may also be used when the data support a nonlinear mode of action, and there is a suggestion of mutagenicity, but the data justifies the conclusion that mutagenicity is not operative at low doses. The guidelines present criteria (based on a modification of the Hill criteria for evaluation of epidemiology data) for evaluation of potential modes of action. When the nonlinear extrapolation is used, an RfD- or RfC-like value is derived using standard methods. Mode of action analysis is critical to the 2005 draft guidelines. This emphasis will bring new research on carcinogenic processes to bear in assessments.

 

The 2005 guidelines also include supplemental guidance for assessing susceptibility from early-life exposure to carcinogens. This guidance states that particular attention should be paid to the potential for higher potency from early-life exposure. Mode of action data should also be evaluated for age-specific differences. If chemical-specific data are available to evaluate the age-specific potency, those data should be used. If chemical-specific data are not identified, but the chemical acts via a mutagenic mode of action, age-dependent adjustment factors are used.

 

References

ATSDR.  1993. ATSDR Cancer Policy Framework.  U.S. Department of Health and Human Services. January.  Available at http://www.atsdr.cdc.gov/cancer.html

 

Health Canada.  1994. Human Health Risk Assessment for Priority Substances.  Environmental Health Directorate. Canadian Environmental Protection Act.  Health Canada, Ottawa, 1994.

 

NSF/ANSI Standard 60.  2009.  Drinking Water Treatment Chemicals - Health Effects.  NSF International, Ann Arbor, MI.  Available for a fee at http://www.techstreet.com/standards/NSF/60_2009?product_id=1628365

 

NSF/ANSI Standard 61.  2009.  Drinking Water System Components - Health Effects.  NSF International, Ann Arbor, MI.  Available for a fee at http://www.techstreet.com/standards/NSF/61_2009?product_id=1656646

 

U.S. EPA (Environmental Protection Agency).  2009. Benchmark Dose Software Version 2.1.1.  National Center for Environmental Assessment, Office of Research and Development. Available at http://www.epa.gov/NCEA/bmds/index.html

 

U.S. EPA (Environmental Protection Agency).  2005.  Guidelines for Carcinogen Risk Assessment.  Washington, DC, National Center for Environmental Assessment. EPA/630/P-03/001b.  NCEA-F-0644b.  Available at http://www.epa.gov/cancerguidelines.

 

US. EPA (Environmental Protection Agency).  1999.  Draft Revised Guidelines for Carcinogen Risk Assessment (External Draft, July 1999).  Risk Assessment Forum, Washington, DC. NCEA-F-0644.  Available at http://epa.gov/cancerguidelines/draft-guidelines-carcinogen-ra-1999.htm

 

U.S. EPA (Environmental Protection Agency). 1996. Proposed Guidelines for Carcinogen Risk Assessment.  EPA/600/P-92/003C.  61 Federal Register pp 17960-18011.  April 23, 1996.  Available at http://www.epa.gov/raf/publications/pdfs/propcra_1996.pdf

 

U.S. EPA (Environmental Protection Agency).  1986. Guidelines for Carcinogen Risk Assessment.  Risk Assessment Forum, Washington, DC. EPA/630/R-00/004.  Available at http://epa.gov/cancerguidelines/guidelines-carcinogen-risk-assessment-1986.htm


 

Noncancer Methods

 

Agency for Toxic Substances and Disease Registry ( ATSDR )

The below explanation of derivation of minimal risk levels (MRLs) developed by ATSDR in their toxicological profiles was extracted from the description provided on ATSDR's web site (http://www.atsdr.cdc.gov/mrls/index.html). Additional information is provided in Pohl and Abadin (1995).  Generally, ITER only presents ATSDR's chronic MRLs.  However, there are two situations in which ITER will discuss the intermediate MRL in the synopsis, if ATSDR has not derived a chronic MRL.  ITER will discuss ATSDR's intermediate MRL if none of the organizations have derived a chronic risk value.  In addition, ITER will discuss ATSDR's intermediate MRL if it is based on the same study as another organization's chronic risk value.


Following discussions with scientists within the Department of Health and Human Services (HHS) and the EPA, ATSDR chose to adopt a practice similar to that of the EPA's Reference Dose (RfD) and Reference Concentration (RfC) for deriving substance-specific health guidance levels for non-neoplastic endpoints. An MRL is an estimate of the daily human exposure to a hazardous substance that is likely to be without appreciable risk of adverse noncancer health effects over a specified duration of exposure. These substance-specific estimates, which are intended to serve as screening levels, are used by ATSDR health assessors and other responders to identify contaminants and potential health effects that may be of concern at hazardous waste sites. It is important to note that MRLs are not intended to define clean-up or action levels for ATSDR or other Agencies.


The toxicological profiles include an examination, summary, and interpretation of available toxicological information and epidemiologic evaluations of a hazardous substance. During the development of toxicological profiles, MRLs are derived when ATSDR determines that reliable and sufficient data exist to identify the target organ(s) of effect or the most sensitive health effect(s) for a specific duration for a given route of exposure to the substance. MRLs are based on noncancer health effects only and are not based on a consideration of cancer effects. Inhalation MRLs are exposure concentrations expressed in units of parts per million (ppm) for gases and volatiles, or milligrams per cubic meter (mg/m3) for particles. Oral MRLs are expressed as daily human doses in units of milligrams per kilogram per day (mg/kg/day).

ATSDR uses the no-observed-adverse-effect-level/uncertainty factor approach to derive MRLs for hazardous substances. They are set below levels that, based on current information, might cause adverse health effects in the people most sensitive to such substance-induced effects. MRLs are derived for acute (1-14 days), intermediate (>14-364 days), and chronic (365 days and longer) exposure durations, and for the oral and inhalation routes of exposure. Currently, MRLs for the dermal route of exposure are not derived because ATSDR has not yet identified a method suitable for this route of exposure. MRLs are generally based on the most sensitive substance-induced end point considered to be of relevance to humans. ATSDR does not use serious health effects (such as irreparable damage to the liver or kidneys, or birth defects) as a basis for establishing MRLs. Exposure to a level above the MRL does not mean that adverse health effects will occur.


MRLs are intended to serve as a screening tool to help public health professionals decide where to look more closely. They may also be viewed as a mechanism to identify those hazardous waste sites that are not expected to cause adverse health effects. Most MRLs contain some degree of uncertainty because of the lack of precise toxicological information on the people who might be most sensitive (e.g., infants, elderly, and nutritionally or immunologically compromised) to the effects of hazardous substances. ATSDR uses a conservative (i.e., protective) approach to address these uncertainties consistent with the public health principle of prevention. Although human data are preferred, MRLs often must be based on animal studies because relevant human studies are lacking. In the absence of evidence to the contrary, ATSDR assumes that humans are more sensitive than animals to the effects of hazardous substances and that certain persons may be particularly sensitive. Thus, the resulting MRL may be as much as a hundredfold below levels shown to be nontoxic in laboratory animals. When adequate information is available, physiologically based pharmacokinetic (PBPK) modeling and benchmark dose (BMD) modeling have also been used as an adjunct to the NOAEL/UF approach in deriving MRLs. 


Proposed MRLs undergo a rigorous review process. They are reviewed by the Health Effects/MRL Workgroup within the Division of Toxicology and Environmental Medicine; an expert panel of external peer reviewers; the agency wide MRL Workgroup, with participation from other federal agencies, including EPA; and are submitted for public comment through the toxicological profile public comment period. Each MRL is subject to change as new information becomes available concomitant with updating the toxicological profile of the substance. MRLs in the most recent toxicological profiles supersede previously published levels.

 

ATSDR Contact Person
Dr. Selene Chou
Division of Toxicology and Environmental Medicine
Agency for Toxic Substances and Disease Registry
1600 Clifton Road, Mailstop F62
Atlanta, Georgia 30333
Telephone: 770-488-3357
E-Mail: SChou@cdc.gov

 

 


 

Health Canada

Health Canada has adopted a threshold toxicants approach for substances classified in Groups IV, V, or VI (see Cancer Risk Assessment Methods text for further information on Health Canada cancer classifications). The following is excerpted from Health Canada's "Human Health Risk Assessment for Priority Substances" (1994). Please refer to this text for a more complete discussion.

 

Threshold toxicants are those for which the critical effect is not considered to be cancer or a heritable mutation. Where possible, a dose (or concentration) of a chemical substance that does not produce any (adverse) effect [i.e., "no-observed-(adverse)-effect-level" (NO(A)EL)] for the critical endpoint is identified, usually from toxicological studies involving experimental animals, but sometimes from epidemiological studies of human populations. If a value for the NO(A)EL cannot be ascertained, a lowest-observed-(adverse)-effect-level (LO(A)EL) is used. The nature and severity of the critical effect (and to some extent, the steepness of the dose-response curve) are taken into account in the establishment of the NO(A)EL or LO(A)EL.


An uncertainty factor is applied to the NO(A)EL or LO(A)EL to derive a Tolerable Daily Intake or Tolerable Concentration (TDI or TC), the intake or concentration to which it is believed that a person can be exposed daily over a lifetime without deleterious effect. They are based on non-carcinogenic effects. Short term excursions above these values do not necessarily imply that exposure constitutes an undue risk to health. Ideally, the NO(A)EL is derived from a chronic exposure study involving the most relevant or sensitive species (where possible, determined based on data on species differences in pharmacokinetic parameters or mechanism of action) or on investigations in the most sensitive sub-population (does not include hypersensitive) in which the route of administration is similar to that by which humans are principally exposed. TDIs or TCs are not generally developed on the basis of data from acute or short term studies (unless observed effects in longer term studies are expected to be similar), although they are occasionally based on data from sub-chronic studies in the absence of available information in adequately designed and conducted chronic toxicity studies, in which case an additional factor of uncertainty is included. Exceptionally, another route of exposure may be used where appropriate, incorporating relevant pharmacokinetic data.

 

The uncertainty factor is derived on a case-by-case basis, depending principally on the quality of the database. Generally, a factor of 1 to 10 is used to account for intraspecies variation and interspecies variation (these may be subdivided to address separately kinetic and dynamic differences). An additional factor of 1 to 100 is used to account for inadequacies of the database which include but are not necessarily limited to, lack of adequate data on developmental, chronic or reproductive toxicity, use of a LO(A)EL versus a NO(A)EL and inadequacies of the critical study. An additional uncertainty factor ranging between 1 and 5 may be incorporated where there is sufficient information to indicate a potential for interaction with other chemical substances commonly present in the general environment. Other considerations and possible adjustments might be made for essential substances or severe, irreversible effects. Numerical values of the uncertainty factor normally range from 1 to 10,000.

 

The value of the TDI or TC is compared to the estimated total daily intake of a chemical substance by the various age groups of the population of Canada and, in some cases, certain high exposure sub-groups or to concentrations in relevant environmental media.

An alternative approach, which may be used where data permit, involves estimation of the "benchmark dose", a model-derived estimate of a particular incidence level (e.g., 5%) for the critical effect. More specifically, the benchmark dose is the effective dose (or its lower confidence limit) that produces a certain increase in incidence above control levels. The advantages of the benchmark dose are that it takes into account the slope of the dose-response curve, the size of the study groups and the variability in the data in establishment of the true threshold.

 

Substances classified as "Possible Carcinogenic to Humans" (Group III) are generally assessed in the above manner. Exceptionally, however, in deriving the TDI or TC an additional uncertainty factor (ranging between 1 and 10) may be incorporated to account for the limited evidence of carcinogenicity.

 


 

NSF International

 

NSF International uses the oral reference dose (RfD) methodology as described in Barnes and Dourson (1988), Dourson (1994), and U.S. EPA (2002) for non-cancer risk assessment.  Specific implementation of this methodology is described in Annex A of NSF International/American National Standard 60 "Drinking water treatment chemicals – Health effects," and of NSF International/American National Standard 61 "Drinking water system components – Health effects."  The LED10 (U.S. EPA, 1995), which is modeled using U.S. EPA (2009) benchmark dose methodology, is favored over the NOAEL in selection of the point of departure if dose-response data on the critical effect can be successfully modeled.  Compound specific uncertainty factors are favored over defaults if sufficient data exist for their derivation.  The margin-of-exposure approach may also be used. 

 

 


 

RIVM

An explanation of RIVM's risk assessment methods is available in the following report:

 

Janssen, PJCM and GJA Speijers.  1997.  Guidance on the Derivation of Maximum Permissible Risk Levels for Human Intake of Soil Contaminants.  Report no. 711701006, National Institute of Public Health and the Environment.  Bilthoven, The Netherlands.  January.  Available at http://www.rivm.nl/bibliotheek/rapporten/711701006.pdf or at http://www.rivm.nl/en/  (click on Search, type "711701006", then click on document).

 


TCEQ

An explanation of TCEQ's methods is available in the publication entitled, TCEQ Guidelines to Develop Toxicity Factors” Available at http://www.tceq.texas.gov/publications/rg/rg-442.html .  This document is a technical guide that details the process of developing Effects Screening Levels (ESLs), Reference Values (ReVs), and Unit Risk Factors (URFs).

 


 

U.S. EPA

 

Noncancer Methods - EPA

EPA defines the oral reference dose (RfD) as “an estimate (with uncertainty spanning perhaps an order of magnitude) of a daily oral exposure to the human population (including sensitive subgroups) that is likely to be without an appreciable risk of deleterious effects during a lifetime.” The inhalation reference concentration (RfC) is similarly defined as “an estimate (with uncertainty spanning perhaps an order of magnitude) of a continuous inhalation exposure to the human population (including sensitive subgroups) that is likely to be without an appreciable risk of deleterious effects during a lifetime.” The estimation of RfDs and RfCs lies squarely in the area of hazard identification and dose response assessment as defined by the National Academy of Sciences (NAS, 1983) report on risk assessment in the federal government.

 

The oral RfD and inhalation RfC are useful reference points for gauging the potential effects of other doses. Doses at the RfD (or less) or concentrations at the RfC (or less) are not likely to be associated with any health risks, and are, therefore, assumed likely to be protective and of little regulatory concern. In contrast, as the amount and frequency of exposures exceeding the RfD or RfC increases, the probability that adverse effects may be observed in a human population also increases. However, it cannot be stated categorically that all doses below the RfD or RfC are acceptable and that all doses in excess of the RfD or RfC are unacceptable. The probability of an effect, the percentage of people affected, and the severity of the risk usually increases as the oral dose or inhalation concentration increases. Therefore, small exceedances of the RfD or RfC will generally result in risk to only the most sensitive individuals in the population, and larger exceedances are generally required before most people are affected. In addition, while exposures at or below the RfD or RfC are protective for sensitive people for most chemicals, such exposures may carry some risk for a sensitive individual for some chemicals. Moreover, the precision of the RfD or RfC depends in part on the overall magnitude of the composite uncertainty and modifying factors used in its calculation. The precision at best is probably one significant figure and more generally an order of magnitude, base 10. As the magnitude of this composite factor increases, the estimate becomes even less precise.

 

The basic assumption in the development of an RfD or RfC is that a threshold exists in the dose rate at or above which an adverse effect will be evoked in an organism. EPA and others consider this assumption to be well-founded. I t is supported by known mechanisms of toxicity of many compounds, which show that a known physiologic reserve must be depleted and/or the repair capacity of the organism must be overcome before toxicity occurs (Klaassen, 2001).

 

For health effects that are not cancer, the U.S. Environmental Protection Agency (EPA, 2002, 2005) and others first identify the critical effect(s), which is “the first adverse effect, or its known precursor, that occurs to the most sensitive species as the dose rate of an agent increases.” Human toxicity data adequate for use in the estimation of RfDs or RfCs are seldom available, but if they are available, they are used in the selection of this critical effect. The use of human data has the advantage of avoiding the problems inherent in interspecies extrapolation.

 

After the critical effect(s) has been identified, EPA generally selects from an overall review of the literature an exposure level (e.g., dose rate for oral studies in mg/kg-day, or air concentration for inhalation studies in mg/m3) that represents the highest level tested at which the critical effect(s) was not demonstrated. This level, the No Observed Adverse Effect Level (NOAEL), is the key datum gleaned from the toxicologist's review of the chemical's entire database and is the first component in the estimation of an RfD or RfC. If a NOAEL is not available, the Lowest Observed Adverse Effect Level (LOAEL) is used. It is not considered appropriate, however, to derive an RfD or RfC from a frank effect, such as lethality. As an alternative to the NOAEL or LOAEL, a benchmark dose (BMD) may be used in this part of the assessment. Advantages and disadvantages of NOAELs and BMDs are described elsewhere (U.S. EPA, 1995).

 

In the absence of appropriate human data, animal data are closely scrutinized. Presented with data from several animal studies, EPA and others first seek to identify the animal model that is most relevant to humans, based on the most defensible biological rationale, for instance using comparative pharmacokinetic data. In the absence of a clearly most relevant species, however, EPA and others generally choose the critical study and species that shows an adverse effect at the lowest administered dose. This is based on the assumption that, in the absence of data to the contrary, humans may be as sensitive as the most sensitive experimental animal species.

 

Uncertainty factors (UFs) are reductions in the dose rate or concentration to account for areas of scientific uncertainty inherent in most toxicity databases. The choice of appropriate uncertainty and modifying factors reflects a case-by-case judgment by experts and should account for each of the applicable areas of uncertainty and any nuances in the available data that might change the magnitude of any factor.

 

U.S. EPA has several publications that describe its use of uncertainty factors (UF) in estimating RfDs and RfCs (e.g., Dourson, 1994; EPA, 2002, 2005). EPA considers five areas of uncertainty in developing RfDs and RfCs. The default value for these factors is 10, but factors of 3 (a half-log of 10, rounded to one significant figure), or 1 are routinely used when partial data are available for these areas of uncertainty (Dourson et al., 1996). EPA's UF for intrahuman variability (designated as H) is intended to account for the variation in sensitivity among the members of the human population. EPA's UF for experimental animal to human extrapolation (designated as A) is intended to account for the extrapolation from animal data to the case of humans. Both of these uncertainty factors can be considered to have components of both toxicokinetics and toxicodynamics, and either component can be replaced by data, when available. More information about this approach, termed chemical-specific adjustment factors (CSAFs) is provided in IPCS (2005). EPA is also in the process of developing its own guidelines addressing this approach. As an example of this type of approach, the use of dosimetric adjustments to the experimental animal NOAEL or LOAEL to estimate the Human Equivalent Concentration (HEC) in the development of RfCs addresses much of the kinetic differences between the experimental animal species and humans. Therefore, a 3-fold, rather than a 10-fold factor is used. EPA's subchronic-to-chronic UF (designated as S) is intended to account for extrapolating from NOAELs or LOAELs identified from less than chronic exposure to chronic levels. EPA's UF for LOAEL-to-NOAEL extrapolation (designated as L) is applied when an appropriate NOAEL is not available to serve as the basis for a risk estimate, and extrapolation from an experimental LOAEL is necessary. An uncertainty factor of 3 is typically used when extrapolating from a minimal LOAEL. EPA's database completeness (designated as D) is intended to account for the inability of any single study to adequately address all possible adverse outcomes (Dourson, 1994; EPA, 2002, 2005).

 

Older EPA assessments occasionally also used an additional factor, referred to as a modifying factor (MF), as an occasional, necessary adjustment in the estimation of an RfC or RfD to account for areas of uncertainty not explicitly addressed by the usual factors. The value of the MF is greater than zero and <10, but it should generally be developed on a log 10 basis (i.e., 0.3, 1, 3, 10) since its precision is not expected to be any greater than the standard UFs. The default value for this factor is 1. Current EPA assessments consider the issues addressed in the context of the MF to fall under the five standard UFs, typically under the database UF.

 

The RfD is composed of the NOAEL or LOAEL or BMD divided by the composite UF, calculated as the product of all individual UFs (and MF, if relevant). The following equation is used:

 

RfD = NOAEL or LOAEL or BMDL / (UF). 
   
The equation that EPA uses to determine the value of the RfC is:    
RfC  =  NOAEL(HEC) or LOAEL(HEC) or BMCL(HEC) (mg/m3) / (UF)    
where: 
   
NOAEL(HEC) = No Observed Adverse Effect Level-Human Equivalent    
                             Concentration      

 
LOAEL(HEC) = Lowest Observed Adverse Effect Level-Human Equivalent    
                            Concentration      

 
 
BMCL(HEC) = Benchmark Concentration Lower Limit -Human Equivalent    
                           Concentration    
 

The "Human Equivalent Concentration" designation reflects the incorporation of dosimetric considerations in the development of an RfC. In determining the dosimetric adjustments between the experimental animal specie and humans, one first determines whether the agent was a particle or a gas (vapor). The approach for dosimetric adjustments for gases is determined by the gas category. Gases are categorized by their target effects and the chemical/physical properties (all of which relate to the mode of action). For particles, the dosimetric adjustments take into account the differences in deposition to different regions of the respiratory tract in the experimental animal specie and humans. These differences depend on the particle size, inhalation rate, and respiratory tract dimensions. Dosimetric adjustments are also available to account for differences between occupational and continuous general population exposures. For a comprehensive understanding of this method the interested reader is referred to U.S. EPA (1994), Jarabek (1994), or Jarabek (1995).

 

Finally, EPA (2005) provides a statement of confidence in its noncancer risk estimates for each chemical on its Integrated Risk Information System (IRIS). High confidence indicates a judgment that additional toxicity data are not likely to change the RfC or RfD (Barnes and Dourson, 1988). Low confidence for an RfD indicates that at least a single, well-conducted, subchronic mammalian bioassay by the appropriate route is available. A low confidence RfC means that at least a single, well-conducted, subchronic mammalian bioassay that identified a NOAEL and included evaluation of the respiratory tract is available. For such a minimum database, the likelihood that additional toxicity data may change the RfC or RfD is greater. Medium confidence indicates a judgment somewhere between these former two choices.

Additional information on methods for developing RfDs and RfCs is provided in U.S. EPA (2002).


References

 

Barnes, D.G., and M.L. Dourson. 1988. Reference Dose (RfD): Description and Use in Health Risk Assessments. Regulatory Toxicology and Pharmacology, 8:471-486.

 

Klaassen, C, Ed. 2001. Casarett and Doull's Toxicology: The Basic Science of Poisons. McGraw-Hill, Medical Publishing Division, New York, NY. pp.  64-78; 92-93.

 

Dourson, M.L., 1994. Methodology for establishing oral reference doses (RfDs). In: Risk Assessment of Essential Elements. W. Mertz, C.O. Abernathy, and S.S. Olin (editors), ILSI Press Washington, D.C., pages 51-61.

 

Dourson, M.L., S.P. Felter and D. Robinson. 1996. Evolution of science-based uncertainty factors in noncancer risk assessment. Reg. Tox. Pharmacol., 24: 108-120.  Available at http://www.tera.org/Publications/UF%20in%20Noncancer.pdf

 

Health Canada. 1994. Human Health Risk Assessment for Priority Substances. Environmental Health Directorate. Canadian Environmental Protection Act. Health Canada, Ottawa, 1994.

 

IPCS (International Programme on Chemical Safety). 2005.  Final Guidance Document for the Use of Data in Development of Chemical Specific Adjustment Factors (CSAFs) for Interspecies Differences and Human Variability:  Guidance Document for Use of Data in Dose/Concentration- Response Assessment, (Harmonization Project Document 2), World Health Organization, Geneva.  Available at http://www.who.int/ipcs/methods/harmonization/areas/uncertainty/en/index.html

 

Jarabek, A.M. 1994. Inhalation RfC methodology: Dosimetric adjustments and dose-response estimation of noncancer toxicity in the upper respiratory tract. Inhal. Tocicol. 6(suppl):301-325.

 

Jarabek, A.M.  1995.  Interspecies extrapolation based on mechanistic determinants of chemical disposition.  Human and Ecological Risk Assessment.  1(5):641-662.

 

NAS (National Academy of Sciences). 1983. Risk Assessment in the Federal Government: Managing the Process. National Academy Press, Washington, DC.

 

NSF/ANSI Standard 60.  2009.  Drinking Water Treatment Chemicals - Health Effects.  NSF International, Ann Arbor, MI.  Available for a fee at http://www.techstreet.com/standards/NSF/60_2009?product_id=1628365

 

NSF/ANSI Standard 61.  2009.  Drinking Water System Components - Health Effects.  NSF International, Ann Arbor, MI.  Available for a fee at http://www.techstreet.com/standards/NSF/61_2009?product_id=1656646

 

Pohl, H.R. and H.G. Abadin. 1995. Utilizing uncertainty factors in minimal risk levels derivation. Regulatory Toxicology and Pharmacology. 22:180-188.


U.S. EPA (Environmental Protection Agency).  2009. Benchmark Dose Software Version 2.1.1.  National Center for Environmental Assessment, Office of Research and Development. Available at http://www.epa.gov/NCEA/bmds/index.html


U.S. EPA (Environmental Protection Agency).  2005.  Integrated Risk Information System (IRIS).  Available at http://www.epa.gov/iris.  IRIS guidance documents and individual chemical files.


U.S. EPA (Environmental Protection Agency).  2002.  A Review of the Reference Dose and Reference Concentration Processes. U.S. EPA, Risk Assessment Forum, Washington, DC, EPA/630/P-02/002F, 2002.  Available at http://www.epa.gov/raf/publications/pdfs/rfd-final.pdf


U.S. EPA (Environmental Protection Agency). 1995. The use of the benchmark dose approach in health risk assessment. Risk Assessment Forum. Office of Research and Development. Washington, D.C. EPA/630/R-94/007.


U.S. EPA (Environmental Protection Agency). 1994. Methods for Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry. Office of Health and Environmental Assessment. Washington, DC. EPA/600/8-90-066F, October.