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Citation: 45 Fed. Reg. 79347 1980

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-- The search text of this PDF is generated from uncorrected OCR text.Federal Register / Vol. 45, No. 231 / Friday, November 28, 1980 / Notices

Residue Value

- (axim m permissible tissue concentration) (mean normalized

a. Foran FDA action level for fish oil, the appropriate percent lipids value is 100.

b. Foran FDA action level for fish, the appropriate percent lipids value is 15 for freshwater criteria and 16 for saltwater criteria because FDA action levels are applied on a species-by-species basis to commonly consumed species. The edible portion ofthe freshwater lake trout averages about 15percent lipids, and the edible portion ofthe saltwater Atlantic herring averages about 16 percent lipids (Sidwell,V. D., et al. 1974 Composition of the Edible Portion of Raw (Fresh orFrozen) Crustaceans, Finfish, and Mollusks.

.Protein, Fat, Moisture, Ash, Carbohydrate, Energy Value, and Cholesterol. Marine Fisheries Review 36:21-35).

c.For a maximum acceptable dietary intake derived from a chronic feeding study with wildlife, the appropriate percent lipids is the percent lipids ofan aquatic species orgroup ofaquatic species which constitute a major portion ofthe diet ofthe wildlife species.

F. The Final Residue Value is obtained by selecting the lowest of the available residue values. It should be noted thatin many cases the Final Residue Value will not be low enough. Forexample, a residue value calculated from an FDA action level would result in

an average concentration in the edible portion ofa fatty species that is at the action level. On the average halfof the individuals of the species would have concentrations above the FDA action level.Also, the results ofmany chronic feeding studies are concentrations that

cause adverse effects.

X. OtherData

Pertinent information that could not be used in earlier sections maybe available concerning adverse effects on aquatic organisms and theiruses. The
most important ofthese are dataon flavor impairment, reduction in survival, growth, orreproduction, or any other adverse effect thathas been shown to be biologically significant. Especially important are data for species forwhich

no other data are available. Data from behavioral. micorcosm, field, and physiological studies may also be available.

XI. Criterion

A.The criterion consists oftwo concentrations, one that should not be BC)(appropriate per'ent lipids)
exceeded on the average in a 24-hour period and one that shouldnotbe exceeded at any time during the 24-hour period. This two-number criterion is intended to identify water quality conditions that should protect aquatic life and its uses from acute and chronic adverse effects ofboth cumulative and noncumulative substances without being

as restrictive as a one-number criterion would have to be toprovide the same degree ofprotection.

B.The maximum concentration is the Final Acute Value oris obtained from the Final Acute Equation.

C. The 2,-hour average concentration is obtained from the Final Chronic Value, the Final Plant Value, and the Final Residue Value by selecting the lowest available value, unless other data (see SectionX) from tests inwhich the toxicant concentrations were measured show that alower value should be used. Iftoxicity is related to a water quality characteristic, the 24-hour average concentration is obtained from the Final Chronic Equation, the Final Plant Value, and the Final Residue Value by selecting the one that results in the lowest concentrations in the normal range of the water quality characteristic, unless other data (see Section X)from
tests inwhich the toxicant concentrations were measured show that a lower value should be used.

D.The criterion is (the 24-hour average concentration) as a 24-hour average and the concentration should
not exceed (the maximum concentration) at any time.

XII.Review

A. On the basis of all available pertinent laboratory and field information, determine ifthe criterion is consistent with sound scientific evidence. Ifit is not, another criterion, either higher or lower, shouldbe derived using appropriate modifications ofthe Guidelines.

These Guidelines were written by Charles E. Stephan, Donald I.Mount, David 1.Hansen,John H.Gentile, Gary A. Chapman andWilliam A. Brungs of the U.S.E.P.A. Environmental Research Laboratories in Corvallis, Oregon, Duluth, Minnesota, GulfBreeze, Florida, and Narragansett, Rhode Island. Numerous otherpeople, many ofwhom do not work for U.S.E.P.A., provided assistance and suggestions. Appendix C-Guidelines and Methodology Used in thePreparation of Health EffectAssessment Chapters of the Consent Decree WaterCriteria Documents

L Objective

The objective ofthehealth effect assessment chapters ofthe ambient water criteria documents is to estimate ambient water concentrations which do not represent a significant riskto the public. These assessments should constitute a review ofall relevant information on individual chemicals or chemical classes in orderto derive criteria thatrepresent, in the case of suspect orproven carcinogens, various levels ofincremental cancerrisk, or,in the case of otherpollutants, estimates of no-effect levels.

Ideally, ambient water quality criteria should represent levels for compounds in ambient water that do notpose a hazard to the human population. However, in any realistic assessment of human health hazard, a fundamental distinction must be made between absolute safety and the recognition of

some risk. Criteria for absolute safety would have to be based on detailed knowledge ofdose-response relationships in humans,including all

sources ofchemical exposure, the types of toxic effects elicited, the existence of thresholds for the toxic effects,the significance oftoxicant interactions, and the variances ofsensitivities and exposure levels within the human population. Inpractice, such absolute criteria cannot be established because ofdeficiencies inboth the available data and the means ofinterpreting this information. Consequently, the individual human health effects chapters propose criteria which minimize or specify thepotential risk ofadverse human effects due to substances in ambient water. Potential social or economic costs and benefits are not considered in the formulation ofthe criteria.

1I. Types of Criteria
Ambient water quality criteria are based on three types ofbiological endpoints: carcinogenicity, toxicity (i.e., all adverse effects otherthan cancer), and organoleptic effects.

For the purpose ofderiving ambient waterquality criteria, carcinogenicity is regarded as anon-threshold phenomenon. Using this assumption, "safe" or "no effect" levels for carcinogens cannot be established because even extremely small doses mustbe assumed to elicit a finite Increase in the incidence ofthe response. Consequently, water quality 

### 79347Federal Register / Vol. 45, No. 231 / Friday, November 28, 1980 / Notices

criteria for carcinogens arepresented as
a range ofpollutant concentrations associated with corresponding incremental risks.

For compounds which do notmanifest any apparent carcinogenic effect, the threshold assumption is used in deriving

a criterion. This assumption is based on the premise that aphysiological reserve capacity exists within the organism which is thought to be depleted before clinical disease ensues. Alternatively, it maybe assumed that the rate of damage willbe insignificant over the life span of the organism. Thus, ambient water quality criteria.are derived fornoncarcinogenic chemicals, andpresumably resultin no observable-adverse-effect levels (NOAELs) in the exposed human population.

In some instances, criteria arebased

on organoleptic characteristics, i.e., thresholds for taste orodor. Such criteria are established when -insufficient information is available on toxicologic effects orwhen the estimate of the level ofthepollutant in ambient water basedon organoleptic effects is lower than the level calculated from toxicologic data. It should be recognized that criteria based solely on organoleptic effects do not necessarily represent approximations ofacceptable risklevels forhuman health.

Several ambient water quality criteria documents deal with classes of compounds which include chemicals exhibiting varying degrees ofstructural similarity. Because prediction of biological effects based solely on structural parameters is difficult, the derivation ofcompound-specific criteria ispreferable to a class criterion.A compound-specific criterion is defined

as a level derived from data on each individual subject compound that does not represent a significant riskto the public. For some chemical classes, however, a compound-specific criterion cannot be derived for each member ofa class. In'such instances, itis sometimes justifiable to derive a class criterion in which available data on one member of

a classmaybe used to estimate criferia for other chemicals ofthe classbecause

a sufficient data base is not available for those compounds.

Forsome chemicals and chemical classes, the data base wasjudged to be insufficient for the derivation of a criterion. In those cases, deficiencies in the available information are detailed.
II. Approach

The human health effects chapters attempt to summarize all information on the individual chemicals or classes of chemicals which mightbe useful in the risk assessmentprocess to develop

- water quality criteria. Although primary emphasis is placed on identifying epidemiologic and toxicologic data', these assessments typically contain discussions on four topics: existing levels ofhuman exposure, pharmacokinetics, toxic effects, and criterion formulation.

For all documents, an attempt is made
to include the known relevant information. Review articles and reports

are often used in the process ofdata evaluation and synthesis. Scientific judgmefit is exercised in thereview and evaluation ofthe data in each document' and in the identification ofthe adverse effects against whichprotective criteria

are sought. In addition, each ofthese documents is reviewed by a peer committee ofscientists familiar with the specific compound(s). These work. groups evaluate the quality ofthe available data, the completeness ofthe data summary, and the validity ofthe derived criterion.

In the analysis and organization ofthe data, an attempt is made to be consistentwithrespect to the format and the application ofacceptable scientific principles. Evaluation procedures used in the hazard
assessment process follow the principles outlined by the National Academy of Sciences in DrinkingWaterandHealth (1977) and the guidelines ofthe Carcinogen Assessment Group ofthe U.S. EPA.

A. Exposure

The exposure section ofthe health effects chapters reviews known information on current levels ofhuman
exposure to the individual pollutant from all sources. Much ofthe data was obtained from monitoring studies ofair, water,food, soil, and human or animal tissue residues. The major purpose of this-section is to-provide background information on the contribution ofwater exposure relative to all other sources. Consequently, the exposure section includes subsections reviewing different
routes ofexposure including water and food ingestion, inhalation, and dermal
contact.

Information on exposure can be valuable in developing and assessing a water quality criterion. In these documents exposure from ionsumpflon ofcontaminated water and contaminated fish and shellfish products is used in criterion formulation. Data for all modes of exposure are useful in relating total intake to the expected contribution from contaminated water, fish, and shellfish. In addition, information for all routes of exposure, not limited to drinking water and fish and shellfish ingestion, can be used to justify or assess the feasibility oftie formulation ofcriteria for ambient water.

The use offish consumption as an exposure factor requires the quantitation ofpollutant residues in the edible portions of the ingested species. Accordingly, bioconcentratlon factors (BCFs)are used torelate pollutant residues in aquatic organisms to the pollutant concentration in the ambient waters inwhich they reside.

To estimate the average per capita intake of apollutant due to consumption ofcontaminated fish and shellfish the results ofa diet survey were analyzed to calculate the average consumption of freshwater and estuarine fish and shellfish (U.S. EPA, 1980). A species is considered to be a consumed freshwater

or estuarine fish-and shellfish species If at some stage in its life cycle, it is harvested from fresh orestuarine water forhuman consumption in significant quantities (Stephan, 1980).

Three different procedures are used to estimate the iveighted average BCF depending upon the lipid solubility of the chemical and the availability of bioconcentration data.

For lipid-soluble compounds, the average BCFis calculated from the weighted average percent lipids in the edible portions ofconsumed freshwater and estuarine fish and shellfish which

was calculated from data on consumption ofeach species and its corresponding percent lipids to be 3.0 percent (Stephan, 1980). Because the steady-state BCFs for lipid-soluble compounds are proportional to percent lipids, bioconcentration factors for fish and shellfish can be adjusted to the average percent lipids for aquatic organisms consumed by Americans. For many lipid-soluble pollutants, there exists at least one BCF forwhich the percent lipid value was measured for the tissues for which the BCF is determined.

With 3.0 percent as the weighted average percent lipids for freshwater and estuarine fish and shellfish in the average diet, a BCF, and a corresponding percent lipid value, the weighted average bioconcentration factor can be calculated.

Example: Weighted average percent lipilds for

average diet=3.0 percent Measured BCF of17 for

trichloroethylen6 with bluegills at 4.8 percent lipids Weighted average BCF for average

diet equals



### 17 

## X = 0. 6 4.8% 

### 79348'Federal Register / Vol. 45, No. 231 

## / Friday, November 28, 1980 

## / Notices

As an estimate, 10.6 is usedfor the

BCF. In those cases where an appropriate bioconcentration factor is not available, the equation "Log BCF= (0.85 Log P)- 0.70" can be used (Veith, et al. 1979] to estimate the BCF for aquatic organisms containing about 7.6 percentlipids (Veith, 1980] from the octanol/water partition coefficient P.An adjustment for percent lipids in the average diet

versus 7.6 percent is made in order to derive the weighted average bioconcentration factor.

Fornon-lipid-soluble compounds, the available BCFs for the edible portion of consumed freshwater and estuarine fish and shellfish are weighted according to consumption factors to determine a weighted BCF representative ofthe
average diet.

B.Pharmacokinetics

This section summarizes the available information on the absorption, distribution, metabolism, and elimination ofthe compound(s) in humans and experimental mammals. Conceptually, such information is useful in validation ofinter- and intraspecies extrapolations, and in characterizing the modes oftoxic action. Sufficient information on absorption and excretion in animals, together with a knowledge of ambient concentrations in water, food, and air, could be useful in estimating body burdens ofchemicals in the human population. Distribution data which suggest target organs or tissues are desirable for interspecies comparison techniques. In terms ofthe derivation of criteria, pharmacokinetic data are essential to estimate equivalent oral doses based on data from inhalation or other routes ofexposure.
C.Effects

This section summarizes information

on biological effects inboth humans and experimental mammals resulting in: acute, subacute, and chronic toxicity, synergism and/or antagonism, teratogenicity, mutagenicity, or carcinogenicity.

The major goal ofthis section is to survey the suitability ofthe data foruse in assessment ofhazard and to determine which biological end-point, i.e., non-threshold, threshold, or organoleptic, should be selected for use in criterion formulation.

Because this section attempts to

assess potential human health effects, data on documented human effects are thoroughly evaluated. However, several factors inherent in human epidemiological studies usually preclude the use ofsuch data ingenerating water quality criteria. These problems, as summarized by the National Academy ofSciences (NAS, 1977) are as follows:

1. Epidemiology cannot tell what effects a material willhave until after humans have been exposed. One must not conduct what might be hazardous experiments on man.

2. Ifexposure has been ubiquitous, it may be impossible to assess the effects of a material, because there is no unexposed control group. Statistics of morbidity obtained before use ofa new material can sometimes be useful, but when latent periods are variable and times ofintroduction and removal of materials overlap, historical data on chronic effects areusually unsatisfactory.

3. It is usually difficult to determine doses in human exposures.

4. Usually,it is hard to identify small changes incommon effects, which may nonetheless be important ifthe population is large.

5. Interactions in a "nature-designed" experiment usually cannot be controlled.

Although these problems often prevent the use ofepidemiological data in quantitative risk assessments, qualitative similarities or differences between documented effects in humans and observed effects in experimental mammals are extremely useful in testing the validity ofanimal-to-man extrapolations. Consequently, in each
case, an attempt is made to identify and utilize both epidemiologic and animal dose-response data. Criteria derived from such a confirmed data base are considered to be reliable.

The decision to establish a criterion based on anon-threshold model is made afterevaluating all available information on carcinogenicity and supportive information on mutagenicity. The approach and conditions for the qualitative decision ofcarcinogenicity

are outlined in the U.S. EPA Interim Cancer Guidelines (41 FR 21402), in a report by Albert, et al. (1977), and in the Interagency Regulatory Liaison Group (IRLG) guidelines on carcinogenic risks (IRLG, 1979). Itis assumed that a substance which induces a statistically significant carcinogenic response in animals has the capacity to cause

cancer inhumans. A chemical which has not induced a significant cancer response in humans orexperimental animals is not identified as a carcinogen, even though Itsmetabolites

or close structural analogues might induce a carcinogenic response orit was shown tobe mutagenic inan in vitro system.

It is recognized that some potential human carcinogens may notbe Identified by the guidelines given above. For example, compounds forwhich there is plausible but weak qualitative evidence ofcarcinogenicity in experimental animal systems (suchas data frommouse skin painting or strain Amouse pulmonary adenoma) would be included in this category. The derivation ofa criterion forhuman consumption from these studies in notvalid, regardless ofthe qualitative outcome. In addition, there are certain compounds (e.g., nickel and beryllium) which were shown tobe carcinogenic inhumans afterinhalation exposure by chemical form,but have inducedthus far no response in animals orhumans via ingesting their soluble salts. Nevertheless, a non-threshold criterion is developed forberyllium because
tumors have been produced in animals
at a site removed from the site of administration; in contrast, a threshold criterion isrecommended fornickel because there is no evidence oftumors at sites distantresulting from administration ofnickel solutionsby either ingestion orinjection.

For those compounds which were not reported to induce carcinogenic effects

orfor those compounds for which carcinogenic data are lacking or insufficient, an attempt is made to estimate a no-effect level. Inmany respects, the hazard evaluation from these studies is similarto that of bioassays forcarcinogenicity. In order
to more closely approximate conditions ofhuman exposure, preference is given
to chronic studies involving oral exposures inwater or diet over a significant portion ofthe animal life span. Greatest confidence is placed in those studies which demonstrate doserelated adverse effects as well as noeffect levels.

There is considerable variability in thebiological endpoints used to define a no-effect level. Theymay range from gross effects, such as mortality, to more subtle biochemical, physiological, or pathological changes. Teratogenicity, reproductive impairment, and behavioral effects are significant toxic consequences ofenvironmental contamination. In instances where carcinogenic or other chronic effects

occur at exposure levels below those causing teratogenicity, reproductive impairment, or behavioral effects, the former are used in deriving the criterion. For most ofthe compounds evaluated thus far, teratogenicity andreproductive impairment occur at dosesnear maximum tolerated levels with dose administration schedules well above estimated environmental exposure levels.Moreover, information on behavioral effects, which could be of I I II I I 

### 79349Federal Register / Vol. 45, No. 231 / Friday, November 28, 1980 / Notices

significance, is not available f6rmost of the compounds under study. Consequently, most NOAELs derived

- from chronic studies are based either on

gross toxic effects or on effects directly related to functional impairment or defined pathological lesions.

For compounds on which adequate chronic toxicity studies are not available, studies on acute and subacute toxicity assume greater significance. Acute toxicity studies usually involve single exposures at lethal or near lethal doses. Subacute studies often involve
exposures exceeding 10 percent ofthe life span ofthe test organism, e.g., 90 days for the rat with an average life
span of30 months. Such stildies are useful in establishing the nature ofthe compound's toxic effects and other parameters ofcompound toxicity, such

as target organ effects, metabolic behavior, physiological/biochemical effects,'and patterns ofretention and tissue distribution. The utility ofacute and subacute studies in deriving environmentally meaningful NOELs is uncertain, although McNamara (1976) has developed application factors for such derivations.

In some cases where adequate data

are not available from studies utilizing oral routes ofadministration, no-effect levels for oral exposures may be estimated from dermal or inhalation studies. Such estimates involve approximations of the total dose administered based on assumptions about breathing rates and/or magnitude ofabsorption.

D. Criterion Rationale

This section reviews existing standaids for the chemical(s), summarizes data on current levels of human exposure, attempts to identify special groups at risk, and defines the basis for the recommended criterion.

Information on existing standards is included primarily for comparison with the proposed water quality criteria. Some of the present standards, such as

-- those recommended by the Occupational Safety and Health Administration (OSHA) or the American Conference ofGovernmental Industrial Hygienists (ACGIH), are based on toxicologic data but are intended as acceptable levels for occupational rather than environmental exposure. Other levels, such as those recommended by the National Academy ofSciences in DrinkingWaterand Health(1977) or in the U.S. EPA Interim' Primary Drinking Water Standards, are

more closely related to proposed water quality criteria. Emphasis is placed on detailing the basis for the existing standards wherever possible. Summaries ofcurrent levels ofhuman exposure, presented in this section, -specifically address the suitability ofthe data to derive water quality criteria. The identification ofspecial groups at risk, either because ofgeographical or occupational differences in exposure or biological differences in susceptibility to the compound(s), focuses on the impact that these groups should have on the development ofwater-quality criteria.

The basis for the recommended criteria section summarizes and qualifies all of.the dataused in developing the criteria.

IV.Guidelines for Criteria Derivation
The derivation ofwater quality criteria from laboratory animal toxicity duta is essentially a two-step procedure: First, a total daily intake forhumans must be estimated which establishes either a definedlevel ofrisk fornonthreshold effects or a no-effect leiel for threshold effects. Secondly, assumptions must bemade about the contribution of contaminated water and the consumption offish/shellfish to the total daily intake ofthe chemical. These estimates are then used to establish the tolerable daily intake and consequently the water quality criterion.

A. Non-ThresholdEffects

After the decision has been made that

a compound has the potential for causing cdncers in humans and that data exist which permit the derivation ofa criterion, the water concentration which is estimated to cause alifetime carcinogenic risk of10- is-determined. Thblifetime carcinogenicity riskis the probability that a person wohld get

cancer sometime inhis orherlife assuming continuous exposure to the compound. The water concentration is calculated by using the low-dose extrapolation procedure proposed by Crump (1980). Thisprocedure is an improvement on the multistage low dose extrapolation procedure by Crump, et al. (1977).

The data usedfor quantitative estimates are oftwo types: (1)lifetime animal studies, and (2)human studies where excess cancer riskhas-been associated with exposure to the agent. In animal studies it is assumed, unless. evidence exists to the contrary, thatif a carcinogenic response occurs at the dose levels used in the study,then proportionately lower responses will also occur at all lower doses, with an incidence determined by the extrapolation model discussed below.

1. Choice ofModel. There is noreally solid scientific basis for any mathematical extrapolation model which relates carcinogen exposure to cancer risks at the -extremely low levels ofconcentration that must be dealt with in evaluating the environmental hazards. For practical
reasons, such low levels ofrisk cannot be measured directly either using animal experiments or epidemiologic studies. We must, therefore, depend on our current understanding ofthe mechanisms ofcarcinogenesis for guidance as to which risk model to use. At the present time, the dominant view ofthe carcinogenic process involves the conceptthat most agents which cause
cancer also cause irreversible damage to DNA. This position Is reflected by the fact that a very large proportion of agents which cause cancer are also mutagenic. There is reason to expect that the quantal type ofbiological response thatis characteristic of mutagenesis is associated with a linear non-threshold dose-response relationship. Indeed, there is substantial evidence frommutagenesis studies with both ionizing radiation and with a wide variety ofchemicals that this type of' dose-response model is the appropriate

one to use. This is particularly true at the lower end ofthe dose-response
curve; at higher doses, there can be an upward curvature, probably reflecting the effects ofmultistage processes on the mutagenic response. The linear nonthreshold dose-response relationship Is also consistent with the relatively few epidemiological studies ofcancer responses to specific agents that contain enough information to make the evaluation possible (e.g., radiationinduced leukemia, breast and thyroid
cancer, skin cancer induced by arsenic in drinking water, and liver cancer induced by aflatoxin in the diet). There is also some evidence from animal experiments thatis consistent with the linear non-threshold hypothesis (e.g., liver tumors induced in mice by 2- .acetylaminofluorene in the large scale EDo 1 study at the National Center of Toxicological Research, and the initiation stage ofthe two-stage carcinogenesis model in the ratliver and the mouse skin).

Because it has the best, albeit limited, scientific basis ofany of the current mathematical extrapolation models, the linear non-threshold model has been adopted as the primary basis forrisk extrapolation tolow levels of the doseresponse relationship. Therisk
assessments made with this model should be regarded as conservative, representing the mostplausible upper limit for therisk; i.e., the true riskis not likely to be higher than the estimate, but it could be smaller. I 79350Federal Register / VoL 45, No. 231 / Friday, November 28, 1980 / Notices

The mathematical formulation chosen
to describe the linear, non-threshold dose-response relationship atlow doses is the improved multistage model developed by Crump (1980).This model employs enough arbitrary constants to be able to fit almost any monotonically increasing dose-response data and it incorporates a procedure for estimating the largestpossible linear slope (in the 95 percent confidence limit sense) at low extrapolated doses that is consistent with the data at all dose levels of the experiment.For thisreason, it may be called a 'linearized" multistage model.

2. Procedure ofLow-Dose Extrapolation Based on Animal Carcinogenicity Data.

A. Description ofthe Extrapolation Model

LetP(d) represent the lifetime risk (probability) ofcancer at dose d. The multistage modelhas the form P(d)=l-exp [-q.+qd+q4d+... +qkdk)j where:

qi>O,and i=O,1,2_

.. k Equivalently.

A(d)=1-exp [-(qd+qd+... 

### +qkdk] where:



### A(d) = P(d) -P(), 

## 1- P(o)

is the extra risk overbackground rate at dose d.

The point estimate of the coefficients 04, i=O, 1, 2

.... ,k, and consequently the extrarisk function A(d) at anygiven dose d,is calculated by maximizing the likelihood function ofthe data.

The point estimate and the 95 percent upper confidence limit ofthe extra risk A(d) are calculated by using the computer program GLOBAL 79 developed by Crump and Watson (1979). Upper95percent confidence limits on the extra risk and lower 95 percent confidence limits on the dose producing

a given risk are determined from a 95 percentupper confidence limit, q1 *,on parameter qi. Whenever q,=*0, at low doses extra riskA(d) has approximately the formA(d)=q

, xd.Therefore, qxd is a 95 percent upper confidence limit on the extra risk and R/qi* isa 95 percent lower confidence limit on the dose producing an extrarisk ofR.Let L. be the maximum value ofthelog-likelihood function. Theupper limit qz* is calculated buy increasing q2 to a value qi* such that when the log-likelihood is again maximized subject to this fixed value q,* for the linear coefficient, the resultingmaximum value ofthe loglikelihood 14 satisfies the equation 2(L.-I,)=2.7G554 where 2.70654 is the cumulative 90 percent point ofthe chi-square distribution with one degree offreedom, which corresponds to a 95percent upper limit (one-sided). This approach of computing the upper confidence limit for the extra riskA(d) is an improvement on the Crump, et al. (1977) model, The upper confidence limit for the extra risk calculated atlow doses is always linear. This is conceptually consistent with the linear nonthreshold concept discussed earlier. The slope q,* is taken as an upper bound of the potency ofthe chemical in inducing cancer at low doses.

In fitting the dose-response model, the number ofterms in the polynomial g Is chosen equal to (h-i]. where his the number ofdose groups in the experiment. including the control group.

Whenever the multistage model does notfit the data sufficiently, data at the highest dose is deleted and the modelis refitted to the rest of the data. This is continued until an acceptable fit to the data is obtained. To determine whether

ornot a fit is acceptable, the chi-square statistic: 

### h

X2= (Xi

- NiPi)2
NiPi (1

- Pi)

is calculated, where N, is the numberof animals in the i'b dose group,X 1 is the number ofanimals in the i" dose group with a tumorresponse, Pi is the probability ofa response in the ill dose group estimatedby fitting themultistage model to the data, and hIs the number ofremaininggroups.

The fit is determined to be unacceptable whenever chi-square (XJ is larger than the cumulative 99 percent point ofthe chi-square distribution with fdegrees offreedom, where fequalsthe number ofdose groups minus the number ofnon-zero multistage coefficients.

3. Selection and Form ofData used to Estimate Parameters in the Extrapolation Model

For some chemicals, several studies in different animal species, strains, and

sexes each conducted at several doses and different routes ofexposure are available. A choice must bemade as to which ofthe data setsfrom several studies are tobe usedin the model. Itis also necessary to correct formetabolism differences between species and for differences in absorption via different
routes ofadministration. The procedures, listed below, used in evaluating these data are consistent with the estimate ofa maximum-likelyrisk. a.The tumorincidence data are separated accordingto organ sites or tumor types.The set data (i.e., dose and
tumorincidence) usedin themodel is setwhere theincidence is statistically significantly higherthan the control for at least one test dose level and/or where thetumor incidence rate shows a statistically significant trend with respect to dose level. The data setwhich gives the highest estimate oflifetime carcinogenic risk q 1*is selected in most

cases. However, efforts are made to exclude data sets which produce spuriously highrisk estimatesbecause ofa small number ofanimals. That is, if two sets of data show a similar doseresponse relationship and one has a very small sample size. the set ofdata which has the larger sample sizeis selectedfor calculating the carcinogenic potency.

b. Ifthere are two ormore data sets of comparable size which are identical with respect to species, strain. sex. and tumorsites, the geometricmean ofqt*, estimated from each ofthese data sets is used for risk assessment. Thegeometric

mean ofnumbersA, Az

... ,A. is defined as (AzxAX

... XA.)J"

c.If sufficient data exist for two or
more significant tumorsites in the same study,the number ofanimals with at least one ofthe specifictumor sites under consideration isused as incidence data in themodel.

d.Following the suggestion ofMantel and Schneiderman (1975). we assume that mg/surface area/day is an equivalent dosebetween species. Since to a close approximation the surface

area isproportional to the %rds power ofthe weight aswould be the case for a perfect sphere, the exposure in mg/%rds power ofthe body weight/day is similarly considered to be an equivalent exposure. In an animal experiment, this equivalent dose Is computed in the following manner.LetL,=duration of experiment 4.=duration ofexposure m=average dose per dayinmgduring administration ofthe agent (Le, during1]) W=averge weightofthe experimental animal Then, thelifetime average exposure is



## d l

## e 

## x m

Le x W 2 

### /3

Often exposures are not giveninunits ofmg/day, and itbecomes necessary to convert the given exposures into mng/ day.For example, in mostfeeding studies, exposure Is expressed as ppmin the diet.In thiscase theexposure (mg/ day) is derived by: m=ppm x F x r 79351Federal Register / Vol. 45, No. 231 / Friday, November 28, 1980 / Notices

where ppm is partsper million ofthe carcinogenic agent in the diet, Fis the weight ofthe food consumedper day in kgms, and r is the absortion fraction.

In the absence ofany data to the contrary, ris-assumed to be one.For a uniform dietthe weight ofthe food consumed is proportional to the calories required, which, in turn, is proportional to the surface' area orthe %rds power of the weight, so that: mappmXW2/ 1 Xr or 

## in ppm

As aresult, ppm in the diet is often assumed to be an equivalent exposure between species. However, we feel that this is notjustified since the calories/kg offood is significantly different in the diet ofman vs. laboratory animals, primarily due tomoisture content differences. Instead, we use an empirically derived food factor, f=F/W, which is the fraction ofa Species bYody weight that is consumed per day as food.We use the rates given below.

Species W f

Man.................. 70 0.028 Rat.

................................... . 0.35 0.05 Mice

...... .................. ... 0.03 0.13

Thus, when the exposure is given as a certain dietary concentration in ppm, the exposure in 4g/W 2 / 3is

m 

## = ppmix 

## F

r x W 2

### /3 

### W2/3

PPM x fxW = 

### ppm x f x W1/3 •W2/3 When exposure is givenin terms of mg/kg/day=nii/Wr=s the conversion is simply:



### m = 

### s x Wl/3 rW 2

### / 3

When exposure is via inhalation, the calculation ofdose canbe considered for two Cases where (1)the carcinogenic agent is either a completely watersoluble gas or an aerosol and is absorbed proportionally to the amount ofair breathed in, and (2) where the carcinogen is a poorly water-soluble gas which reaches an equilibrium between the air breathed and the body compartments. After equilibrium is reached, the rate ofabsorption ofthese agents is expected tobe proportional to metabolic rate, which in turn is proportional to therate ofoxygen consumption, which in turn is a function ofsurface area. Case 1 Agents that are in the form of particulate iatter orvirtually completely absorbed gases such as S02

can reasonably be expected to be absorbed proportional to the breathing rate. In this case the exposure in mg/day may be expressed as: m=IXvxr where I is inhalation rate per day inm, v is mg/m a ofthe agent in air; and r is the. absorption fraction.

The inhalation rates, I,for various species canbe calculated from the observation (FASEB, 1974)that 25 gm mice breathe 34.5 liters/day and 113 gm rats breathe 105 liters/day. For mice and rats of otherweights, W, (expressed in kg], the surface areaproportionality can be used to-determine breathing rates (in ms/day) as follows:

Formice, I=0.0345 (W/0.025) 2/ 1!10/ day

-

Forrats, I=0.105 (W/0.113)21Sms/day Forhumans, the values of20 m3/daya is adopted as a standard breathing rate (ICRP, 1977).

The equivalent exposure inmg/W 2/s
for these agents can be derived from the air intake data in a way analogous to the food intake data. The empirical factors for the air intake perkg per day, i=I/Wbased upon the previously stated relationships, are as tabulated below:

Species W

Man

_ _79 0.29 Rat. 0.35 0.64 Mice

..... 0.03 1.3

Therefore, forparticulates orcompletely absorbed gases, the equivalent exposure inmg/W2'3 is:

mr Ivr = iWvr = jw1/3 vr 

## T

2

## 1_3 72

## 1/3 7213

Inthe absence ofempirical data or a sound theoretical argument to the contrary, the fraction absorbed, r,is assumed to bethe same for all species. Case2

The dose in mg/day ofpartially soluble vapois is proportional to the Qt' consumption whichin turn is proportional toW 21 3and to the solubility ofgas in body fluids, which

canbe expressed as an absorption coefficient r for the gas. Therefore, when expressing the 06consumption as O 2=k W2 /3 ,where kis a constant independent

aFrom "Recommendation of theInternational Commission onRadiological Protection." page 9, the iverage breathing rate is 107cm~per'8-hour work day and 2X0 7 cm3in 24hours. ofspecies, it follows that m=kW21 x V

x r or


### d m kvr 

## w213'

As with Case 1,in the absence of experimental information or a sound theoretical argument to the contrary, the absorption fraction, r,is assumed to be the same for all species. Therefore, for these substances a certain concentration. in ppm or p/m 3in experimental animals is equivalent to the same concentration in humans. This is supported by the observation that the minimum alveolar concentration, necessary to produce a given "stage" ofanesthesia, Is similar in

man and animals (Dripps, et al. 1977). When the animals were exposed via the oral route and human exposure Is via inhalation orvice-versa, the assumption is made, unless there is pharmacokineflo evidence to the contrary, that absorption is equal by either exposure route,

e. Ifthe duration ofexperiment (L.) Is less than the natural life span of the4est animal (L), the slope qa*, ormore generally the exponent g(d), is increased by multiplying afactor (L/L.) 3

. We

assume that ifthe average dose, d,is continued, the age specific rate of

cancer will continue to increase as a constant function ofthebackground rate. The age specific rates forhumans ,increase at leastby the 2ndpower ofthe age and oftenby a considerably higher power, as demonstrated by Doll (1971). Thus, we would expect the cumulative tumorrate to increase by at least the 3rd power of age.Using this fact,we assume that the slope q,*, or more generally, the exponent g(d), would also increase by at least the 3rd power ofage. As a result,if the slope q 1* [or g(d)] is calculated at age L 0,we would expect that ifthe experiment hadbeen continued for the full-life span, L, at the given average exposure, the slope q,* [or g(d)] would have been increased by at least (L/L,)

This adjustment is conceptually consistent to the proportional hazard modelproposed by Cox (1972) and the time-to-tumor model considered by Crump, et al. (1977) vhere the probability ofcancer at age t and dose d is givenby P(d,t)=1-exp[-f(t)Xg(d)]

4.Calculation ofCarcinogenic Potency Based on Human Data.Ifhuman epidemiology studies and sufficiently valid exposure information are available for the compound, they are always used in some way.Ifthey show a carcinogenic effect, the data are analyzed to give an estimate ofthe linear dependence of cancer rates on lifetime average dose, which Is equivalent to the factor q*. Ifthey show 79352Federal Register / VoL 45, No. 231 / Fridak. November 28, 1980 / Notices

no carcinogenic effect when positive animal evidence is available, then it is assumed that arisk does exist but it is smaller than could have been observed in the epidemiologic study, and an upper limit ofthe cancer incidence is calculated assuming hypothetically that the trueincidence isjust below the level ofdetection in the cohort studied, which is determined largely by the cohort size. Whenever possible, human data are used in perference to animal bioassay data.

In human studies, the response is measured in terms of the relative risk of the exposed cohort ofindividuals compared to the control group. In the analysis ofthis data, it is assumed that the excess risk, orrelative risk minus

one, R[X -1, is proportional to the lifetime average exposure, X, and that it is the same for all ages.It follows that the cardnogenic potency is equal to [R(X)-1]/X multiplied by the lifetime risk at that site in the general population. Except for anunusually well-documented human study, the confidence limit forthe excess risk is
not calculated, due to the difficulty in accounting for the uncertainty inherent in the data (exposure and cancer response).

5. Calculation ofWater Quality Criteria. After the value ofqi* in [mg/ kg/day)-I has been determined, the lifetime risk,P, from an average daily
exposure ofx mg/kg/day is foundfrom the equation P=q,*x.Therefore, ifthe lifetime risk is set at P=10-5 for calculation purposes, the intake, I, in mg/day for a 70 kgperson can be found by the equation: I=70X10-S/q* The intake ofthe agent from ambient
water is assumed to come from two

sources: (1) drinking an average of2 liters ofwater per day, and (2) ingesting

an average of6.5 grams offish per day. Because ofaccumulation ofresidues in fish, the amount ofthe pollutant in fish (mg/kg ofedible fish) is equal to a factor R times the water concentration (mg/kg ofwater). Therefore, the total intake I

can be written as sum oftwo terms: I(mg/day) =C(mg/l)XRl/kg fish)x0.0065 kg fish/day+C(mg/lx21/ day=C(2+0.0065R) where C is the
water concentration in mg/L Therefore, the water concentration in mg/l corresponding to a lifetime risk of 10-6 for a 70 kgperson is calculated by the formula:

$C 

### = 70 x 10-5 C = 

### o1*(2 + 0.0065 R) B. Threshold Effects
1.Use ofAnimal Toxicity Data (Oral).

ndeveloping guidelines for deriving criteria based on noncarcinogenic responses, five types ofresponse levels

are considered: NOEL-No-Observed-Effect.Level NOAEL--No-Observed-Adverse.Effect.Level LOEL--Lowest-Observed.Effect.LevelLOAEL--Lowest-Oberved-Adverse-EffectLevel FEL-Frank-Effect-Level

Adverse effects are defined as any effects which result in functional impairment and/or pathological lesions which may affect the performance of the whole organism, orwhich reduce an organism's ability to respond to an additional challenge.

One ofthe major problems encountered in consideration ofthese concepts regards the reporting of "observed effect levels" as contrasted to "observed adverse effect levels". The
terms "adverse" vs. "not adverse" are at times satisfactorily defined, but dueto increasingly sophisticated testing protocols, more subtle responses are being identified, resulting in a needfor judgment regarding the exact definition ofadversity.

The concepts listed above (NOEL, NOAEL, LOEL, LOAEL)have received much attention because they represent landmarks which help to define the threshold region in specific experiments. Thus, Ifa single experiment yields a NOEL a NOAEL'a LOAEL and a clearly defined FELinrelatively closely spaced doses, the threshold region has been relatively well defined. such data

are veryuseful forthe purpose of deriving a criterion. On the otherhand. a clearly defined FELhas littleutility in establishing criteria when it stands alone, because such a level gives no indication how farremoved the data point is from the threshold region. Similarly, a free-standing NOEL has little utility,because there is no indication ofits proximity to the LOEL. since a free-standing NOELmaybe many orders ofmagnitude below the threshold region.

Based on the above dose-response classification system, the following guidelines for deriving criteria have been adopted.

a. A free-standing FELis unsuitable for the derivation ofcriteria.

b.A free-standing NOELis unsuitable for the derivation of criteria. Ifmultiple NOELs are available without additional data on LOELs, NOAELs, orLOAELs, the highest NOEL should beused to derive a criterion.

c.A NOAEL,LOEL, orLOAEL can be suitable forcriteria derivation. A welldefined NOAEL from a chronic (atleast 90-day) studymay be used directly, applying the appropriate uncertainty factor.For a LOEL, ajudgment needs to be made whether it actually corresponds to a NOAEL ora LOAEL Inthe case of

a LOAEL, an additional uncertainty factor Is applied: the magnitude ofthe additional uncertainty factoris judgmental and should lie in the range of 1 to 10. Caution mustbe exercised notto substitute "Frank-Effect-Levels" for"Lowest-Observable-Adverse-EffectLevels".

d. Ifforreasonably closely spaced doses only a NOEL and a LOAEL of equal quality are available, thenthe appropriate uncertainty factor is applied
to the NOEL

In using this approach, the selection andjustification ofuncertainty factors

are critical. The basicdefinition and guidelines for using uncertainty factors has been given by theNational Academy ofSciences (1977). "Safety Factor" or "Uncertainty Factor" is defived as a number that reflects the degree oramount ofuncertainty that mustbe considered when experimental data in animals areextrapolated to man. When the quality and quantity of experimental data are satisfactory, a low uncertainty factoris used;when data isjudged tobe inadequate or equivocal, alarger uncertainty factor is used. The following general guidelines have been adopted in establishing the uncertainty factors:

a. Valid experimental results from studies onprolonged ingestion byman. with no indication ofcarcinogenicity. Uncertainty Factor=10

b.Experimental results ofstudies of human ingestionnot available or scanty (e.g., acute exposure only) with valid results oflong-term feeding studies on experimental animals, or in the absence ofhuman studies, valid animal studies

on one ormore species. No indication of carcinogenicity. Uncertainty Factor=100

c. Nolong-term oracute human data. Scanty results on experimental animals with no indication ofcarcinogenicity. Uncertainty Factor=1,000 Considerable judgment mustbe used in selecting the appropriate safety factors for deriving a criterion. In those cases where the data do not completely fulfill the conditions for one category and appear to be intermediate between two categories an intermediate uncertainty factoris used. Such an intermediate uncertaintyfactor maybe developed based on a logarithmic scale (e.g., 33, beinghalfway between 10 and100 on a logarithmic scale).

In determining the appropriate use of the uncertainty factors, thephrase "no 79353Federal Register / Vol. 45, No. 231 / Friday, November 28, 1980 / Notices

indication ofcarcinogenicity" is interpreted as~the absence of carcinogenicity data from animal experimental studies or human epidemiology. Available short-term carcinogenicity screening tests~are reported in the criteria documents,but they arenot used either for derivation of numerigal criteria nor to rule out the uncertainty factor approach.

Because ofthe high degree of judgment involved in the selection ofa safety factor, the criterion derivation section ofeach document should provide a detailed discussion and justification forboth the selection of the safety factor and the data to which it is applied. This discussion should reflect a critical review ofthe available data base. Factors to be considered include number ofanimals, species, and parameters tested; quality ofcontrols; dose levels; route; and dosing schedules. An effort should be made to differentiate between results which constitute a toxicologically sufficient data base and data which may be spurious in nature.

2. Use ofAcceptable Daily Intake (ADI). Forcarcinogens, the assumption oflow dose linearity precludes the necessity for defining total exposure in the estimation ofincreased incremental risk.For non-carcinogens, ADIs-and criteria derived therefrom are calculated from total exposure data that include contributions from the diet and air.The equation used to derive the criterioh (C) is: C=ADI-(DT+IN)/[2 

### l+(0.0065 kg 

### X R)] where 2 1is assumed daily water consumption, 0.0065 kg is assumed daily fish consumption, R is bioconcentration factor in units of1/kg, DTis estimated non-fish dietary intake, and IN is estimated daily intake by inhalation.

If estimates ofIN and DT cannot be provided from experimental data, an assumption must be made concerning total exposure. Itis recognized that either the inability to estimate DT and IN due tolack ofdata or the wide variability in DT and IN in different' states may add an additional element of uncertainty to the criterion formulation process. In terms ofscientific validity, the accurate .estimate ofthe Acceptable Daily Intake is the major factor in

- satisfactory derivation ofwater quality criteria.

3.Use ofThreshold LimitValues or Animal Inhalation Studies. Threshold Limit Values (TLVs) are established by the American Conference of Governmental and Industrial Hygienists (ACGIH) and represent 8-hour timdweighted average concentrations in air that are intended to protect workers from various adverse health effects over
a normal working lifetime. Similar values are setby NIOSH (criteria) and OSHA (standards) for 10- and 8-hour exposures, respectively. To the extent that these values are based on sound toxicologic assessments and have been protective in the work environment, they provide useful information for deriving

or evaluating water quality criteria. However, eachTLV must be carefully

- examined to determine if the basis of the TLV contains data which can be used directly to derive-awater quality criterion using the uncertainty factor approach. In addition, the history of eachTLV must be examined to assess the extent to which it has assured worker safety. In each case, the types of effects against-which TLVs are designed toprotect are examined in terms of their relevance to exposure from water. It must be demonstrated that the chemical is not a localized irritant and that there is no significant effect at the site of entry irrespective ofthe routes of exposure (i.e., oral or inhalation).

Ifthe TLV or similar value is recommended as thebasis ofthe criterion, consideration ofthe above points is explicitly stated in the criterion derivation section ofthe document. Particular emphasis is placed on the quality ofthe TLV relative to the available toxicity data that normally is givenpriority overTLVs or similar established values. Ifthe TLV can be justified as thebasis for the cirterion, then the problems associated with the estimation.of acceptablezoral doses from inhalation data mustbe addressed.

Estimating equivalencies ofdoseresponse relationships from one route of exposure to another introduces an additional element ofuncertainty in the derivation of criteria. Consequently, whenever possible, ambient water quality criteria should be based on data involving oral exposures. iforal data are insufficient,'data from otherroutes of exposure may be useful in the criterion derivation process.

Inhalation data, including TLVs or similar values, are the most common alternatives to oral data. Estimates of equivalent doses canbe based upon: (1) available pharmacokinetic data for oral and inhalation routes, (2) measurements of absorption efficiency from ingested or inhaled chemicals, or (3] comparative excretion data when the associated metabolic pathways are equivalent to those following oral ingestion or inhalation. Given that sufficient pharmacokinetic data are available, the

use ofaccepted pharmacokinetic models provides the most satisfactory approach for dose conversions. However, if available phiarmacokinetic data are marginal or ofquestionable quality, pharmacokinetic modeling is inappropriate.

The Stokinger and Woodward (1958) approach, or similar models based on assumptions ofbreathing rate and absorption efficiency, represents possible alternatives when data are not sufficient tojustify pharmacokineto modeling. Such alternative approaches, however, provide less satisfactory approximations because they are not based on pharmacokinetic data. Consequently, inusing the Stokinger and Woodward or related models, the" uncertainties inherent in each of the assumptions and the basis ofeach assumption must be clearly stated in the derivation ofthe criterion.

The use ofdata pertaining to other
routes ofexposure to derive water quality criteria may also be considered. As with inhalation data,'an attempt Is made to use accepted toxicologic and pharmacokinetic principles to estimate equivalent oral doses. If simplifying assumptions are used, theirbases and limitations must be clearly specified.

Because ofthe uncertainties involved ,in extrapolating from one route of

exposure to another and the consequent limitations that this may place on the derived criterion, the decision to disallow such extrapolation and recommend no criterion is highly judgmental and must be made on a c'

., by-case basis. A decision for or against criteria derivation must balance the quantity and quality ofthe available data against a perceived riskto the human population.

Ifthe Stokinger and Woodward (1958) approach is used to calculate an ADI from a TLV, the general equation is: ADI=TLVXBRXDEX dXAA/(AoXSF) where: ADI=Acceptable daily intake In mg TLV= Concentration in airin mg/m 3
DE=Duration ofexposure In hours per day d=5 days/7 days AA=Efficiency ofabsorption from air Ao=Efficiency ofabsorption from oral exposure SF=Safety factor following guidelines given above BR=Amount ofair breathed per day; assume 10m3

For deriving an ADI from animal toxicity data, the equation Is: ADI=CAXDExdxAAXBRX70 kg/ (BWA XAoX SF) where: ADI=Acceptable daily intake in mg
CA= Concentration in air In mg/m 3 DE=Duration ofexposure in hours per day d=Number ofdays exposed/number ofdays observed AA=Efficiency ofabsorption from air BR=Volume ofair breathed per day in m 3 70 kg=Assumed human body weight BWA=Body weight ofexperimental animals in kg A354 

# 79a54Federal Register / Vol. 46. No. 231 / Friday, November 28, 1980 / Notices

AO=Efficieay ofabsorption from oral exposure SF=Safety factorfollowing guidelines given above.

More foaln pharmacokinetic models
must be developed on a compound-bycompound basis.

It should be noted that the safety factors used in the above formulae are intended to account for species variability. Consequently. the rng/ surface arealday conversion factor is
not used in the derivation oftoxicity based criterion.

C. OrganolepticCriteria

Organoleptic criteria define concentrations ofmaterials which impart undesirable taste and/or odor to water. In developing and utilizing such criteria two factors must be appreciated: the limitations ofmost organoleptic data and the humanhealth significance of organoleptic properties.

The publications which report taste and odor thresholds are, with very few exceptions, cryptic in their descriptions oftest methodologies, number of subjects tested, concentration: response relationships, and sensory characteristics at specific concentrations above threshold. Thus, the quality of organoleptic data is often significantly less than that oftoxicologic data used in establishing other criteria. Consequently, a critical evaluation of the available organoleptic data must be made and the selection ofthe most appropriate data base for the criterion
must bebased on sound scientific judgment

Organoleptic criteria are not based on toxicologic information and have no direct relationship topotential adverse human health effects. Although sufficiently intense organoleptic characteristics could result in depressed fluid intake which, in turn, might aggravate a variety of functional disease
states (i.e., kidney and circulatory diseases), such effects are not used in the derivation process oforganoleptic criteria unless available data would indicate an indirect human health effect via decreased fluid consumption, criteria derived solely from organoleptic data are based upon aesthetic qualities only.

Since organoleptic and human health effects criteria are based on different endpoints, a distinction must be made between these two sets ofinformation. In criteria summaries involving both types ofdata, the following format is
usech

For comparison purposes, two approaches

were used to derive criterion levels for

.Based on available toxicity data. forthe protection ofpublichealth thederived level is -. Using available organoleptic data. for controlling undesirable taste and odor quality ofambient water the estimated level is -. It shouldberecognized that organoleptic data as a basis for establishing a water quality criteria have no demonstrated relationship to potential adverse human health effects.

In those instances where a level to limit toxicity cannotbe derived, the following statement is to be appropriately inserted:

Sufficient data are not available for

-to derive a level which would protect against thepotential toxicity ofthis compound.

D. Criteria for ChemicalClasses
A chemical class is broadly defined as any group ofchemical compounds which

are reviewed in a single risk assessment document. In criterion derivation, isomers should be regarded as a part of

a chemical class rather than as a single compound. A class criterion is an estimate ofrisk/safety which applies to

more than one member ofa class. It involves the use of available data on

one or more chemicals of a class to derive criteria for other compounds of the same class in the event that there

are insufficient data available to derive compound-specific criteria.

A class criterion usually applies to each member of a class rather than to the sum of the compounds within the class. While the potential hazards of multiple toxicant exposure are not to be minimized, a criterion, by definition.
most often applies to an individual compound. Exceptions may be made for complex mixtures which are produced, released, and toxicologically tested as mixtures (e.g., toxaphene and PCBs). For such exceptions, some attempt is made
to assess the effects ofenvironmental partitioning (i.e., different patterns of environmental transport and degradation) on the validity of the criterion. If these effects cannot be assessed. In appropriate statement of uncertainty should accompany the criterion.

Since relatively minor structural changes within a class of compounds

can have pronounced effects on their biological activities, reliance on class criteria should be minimized. Whenever sufficient toxicologic data are available

on a chemical within a class, a compound-specific criterion should be derived. Nonetheless, for some chemical classes, scientific judgment may suggest

a sufficient degree ofsimilarity among chemicals within a class to justify a class criterion applicable to some ofall members ofa class.

The development ofa class criterion takes into consideration the following: 1.A detailed review ofthe chemical and physical propertiesof chemicals within the group should be made.A close relationship within the class withrespect to chemical acitivity would sugest a similar potential to reach commonbiological sites within tissues. Likewise, similar lipid solubilities would suggest the possibility ofcomparable absorption and tissue distribution.

2. Qualitative and quantitative data for chemicals within thegroup are examined. Adequate toxicologic data ona number of compounds within agroup provides amore reasonable basisfor extrapolation to other chemicals ofthe same class thanminimal data on one chemical ora few chemicals within thegroup.

3.Similarities in thenature of the toxicologic response to chemicals in the class provides additional support forthe prediction that the response to other members ofthe class may besimilar. In contrast where the biological response has been shown to differ markedly on a qualitative and quantitative basis for chemicals within a class,the extrapolation ofa criterion toother members ofthat class is notappropriate.

4.Additional support forthe validity of extrapolation ofa criterion toother members of a class couldbe provided by evidence of similar metabolic and pharmacokinetic data fursome members of theclass.

Based on the above considerajions, it
may be reasonable in some cases to divide a chemical class into various subclasses. Such divisions could be based on biological endpoints (e.g., carcinogens/non-carcinogens), potency, and/or sufficiency ofdata (e.g., a criterion for some members of a class but no criterion for others). While no a priori limits can be placed on the extent ofsubclassification, each subclassification must be explicitly justified by the available data.

Class criteria, ifproperly derived and supported, can constitute valid scientific
assessments ofpotential risk/safety. Conversely, the development ofa class criterion from an insufficient data base

can lead to serious errors in underestimating oroverestimating risk/ safety and should be rigorously avoided. Although scientific judgment has a proper role in the development ofclass criteria, such criteria are uieful and defensible only if they are based on adequate data and scientific reasoning. The definition ofsufficient data on similarities in physical, chemical. pharmacokinetic or toxicologic properties tojustify a class criterion
may vary markedly depending on the degree ofstructural similarity and the gravity ofthe perceived risk. Consequently, it is imperative that the criterion derivation section ofeach document in which a class criterion is recommended explicity address each of the key issues discussed above, and define, as clearly as possible, the 79355Federal Register /Vol. 45, No. 231 

## / Friday, November 28, 1980 / Notices

limitations ofthe proposed criterion as well as the type ofdata needed to generate a compound-specific criterion.

A class criterion should be abandoned when'there is sufficient data availabe to derive a compound-specific criterion which protects against the biological effect ofprimary concern; e.g., the. availability ofa good subchronic study would not necessarily result in the abandonment ofa class criterion based

on potential carcinogenicity.

The inability to derive a valid class criterion does not, and should not, preclude regulation ofa compound or group ofcompounds based on concern forpotential human health effects. The failure to recommend a criterion is simply a statement that the degree of

concern cannot be quantified based on the available data and risk assessment methodology.
E.Essential Elements
Some chemicals, particularly certain metals, are essential to biological organisms at low levels but may be toxic and/ or carcinogenic at high levels. Because ofpotential toxic effects, it is legitimate to establish criteria for such essential elements. However, criteria must consider essentiality and cannot be established at levels which would result in deficiency ofthe element in the human population.

Elements are accepted as essential if listed by NAS Food and Nutrition Board

or a comparably qualified panel. Elements not yet determined tobe essential but for which supportive data

on essentiality exists need to be further reviewed by such a panel.

To modify the toxicity and' carcinogenicity based criteria,: essentiality must be quantified either as
a "recommended daily allowance" (RDA) or "minimum daily requirement" (MDR). These levels are then compared to estimated daily doses associated with the adverse effect ofprimary concern. The difference between the RDA or MDR and the daily doses causing a specified risk level for carcinogens or ADIs fornon-carcinogens defines the spread ofdaily doses from which the criterion maybe derived. Because errors
are inherent in defining both essential and maximum tolerable levels, the criterion is derived from dose levels

near the center ofsuch a dose range. 'The decisiorto use either the MDR or RDA is guided by the spread ofthe doses and the quality ofthe essentiality and toxicity estimates.

The modification of criteria by consideration ofessentiality must take into accountall routes of exposure. If water is a significant source ofthe'MDR

orRDA, the criterion must allow for attainment ofessential intake. Conversely, even when essentiality may be attained from nonwater sources, standard criteria derivation methods maybe adjusted ifthe derived criterion represents a small fraction ofthe ADI or MDR. On a case-by-case basis, the modification in the use ofthe guidelines may include the use of different safety factors for non-carcinogens or other modifications which can be explicitly justified.

F. Useof ExistingStandards
For some chemicals forwhich criteria

are tobe established, drinking water standards already exist. These standards represent not only a critical assessment ofliterature, but also a body ofhuman experience since their promulgation. Therefore, itis valid to accept the existing standard unless there is compelling evidence to the contrary. This decision shouldbe made after considering the existing standards

vs. new scientific evidence which has-

. accumulated since the standards have been established. There are several instances where the peer review process recommended usage ofthe present drinking water standards.

* Bibliography

Albert, R.E., etal. 1977. Rationale developed by the Environmental Protection Agency for the assessment ofcarcinogenic risks. Jour.Natl. Can. Inst. 58:1537.

- Cox, C.R. 1972. Regression model and life tables. Jour. Roy. Stat. Sec. B.34: 187.

Crump, K.S. 1979. Dose-response problems in carcinogenesis. Biometrics. 35: 157.

Crump, K.S. 1980. An improved procedure for low-dose carcinogenic risk assessment from animal data. Jour.Environ. Path. Toxicol. (Inpress).

Crump, K.S. andW.W. Watson. 1979. GLOBAL 79. AFORTRAN program to extrapolate dichotomous animal carcinogenicity data to low dose. NatI. Inst. Environ. Health Sci. Contract No. 1-ES-2123.

Crump, K.S., et al. 1977. Confidence, intervals and test ofhypotheses concerning dose-response relations inferred from animal carcinogenicity data. Biometrics. 33: 437.

Doll, R. 1971. Weibull distribution of

cancer: Implications for models of carcinogenesis. Jour. Roy. Stat. Soc.A. 13: 133.

Dripps, R.D., etal. 1977. Introduction to Anesthesia, the Principles ofSafe Practice. 5th ed. W.B. Saunders Co., Philadelphia, Pennsylvania. p. 121.

Druckrey,H. 1967. Quantitative Aspects of Chemical Carcinogenesis. In:R. Turhaut (ed.), Potential Carcinogenic Hazards from Drugs, Evaluation ofRisks. UICC Monograph Series, Vol. 7. Springer Verlag, New York.

Federation ofAmerican Societies for Experimental Biology. 1974, Biological Data *Books, 2nd ed.Vol. I. P.L.Altman and D.S.

Dittment (eds.)Federation ofAmerican Societies for Experimental Biology. Bethesda, Maryland.Library ofCongress No. 72-87738. Guess, H., etal. 1977. Uncertanlty estimates for low dose rate extrapolations of animal carcinogenicity data. Can, Res. 37: 3475.

Interagency Regulatory Liaison Group. 1979. Scientific bases for identifying potential carcinogens and estimating their risks, February 6.

International Commission on Radiological Protection. 1977. Recommendation of the ICRP, Publ. No. 26, adopted January 17,1077. Pergamon Press, Oxford, England,

Mantel, N. and M.A. Schneiderman, 1975, Estimating "safe lavels," a hazardous undertaking. Can. Res. 35: 1379.

McNamara, B.P. 1976. Concepts In Health Evaluation ofCommercial and Industrial .Chemicals. In:M.A. Mehimar, et al. (eds.), Advances inModem Toxicology, Vol. 1, Part 1. Wiley and Sons, New York, p. 61

National Academy ofSciences. 1977. Drinking Water and Health. National Academy ofSciences, Washington, D.C.

National Cancer Institute. 1970. Guidelines for carcinogen bioassay in small rodents. NCI Carcinogenesis Technical Report Series, No. 1, February, 1976.

Peto, R. 1973. Weibull distribution for continuous-carcinogenesis experimeots. Biometrics. 29: 457.

Stephan, C.E. 1980. Memorandum to J. Stara. U.S. EPA. July 3.

Stokinger, M.E. and R.L. Woodward, 1950. Toxicologic methods for establishing drinking
water standards. Jour. Am. Water Works Assoc. 50: 517.

U.S. EPA. 1976. NationalInterim Primary Drinking Water Regulations. Office of Water Supply. EPA570/9-76-003. U.S. Environ. Prot, Agency.

U.S. EPA. 1980. Seafood consumption data analysis. Stanford Research Institute International, Menlo Park, California, Final report, Task 11, Contract No. 08-01-3887.

Veith, G.D. 1980. Memorandum to C.E. Stephan. U.S. EPA April 14.

Veith, G.D., et al. 1979. Measuring and estimating the bioconcentration factors of chemicals in fish. Jour. Fish Res. Board Can. 36: 1040.

Health Effects Methodology-Authors, Contributorsand Reviewers Jerry F. Stara, ECAO-Cin, U.S. Environmental

Protection Agency Elizabeth L.Anderson, OHEA, U.S.

Environmental Protection Agency Roy E. Albert, N.Y. University Medical

ISchool, New York, NY Patrick R. Durkin, Syracuse Research

Corporation, Syracuse, NY Steven D. Lutkenhoff, ECAO, U.S.

Environmental Protection Agency Robert E. McGaughy, CAG, U.S.

Environmental Protection Agency Charles Ris, OHEA, U.S. Environmental

Protection Agency Lawrence Anderson, CAG, U.S.

Environmental Protection Agency Dolph Arnicar, CAG, U.S. Environmental

Protection Agency Steven Bayard, CAG, U.S. Environmental

Protection Agency Edward Calabrese, University of

Massachusetts, Amherst, MA Thomas Clarkson, University ofRochester,

Rochester, NY 79356Federal Register / Vol. 45. No. 231 / Friday, November 28, 1980 / Notices

Joseph Cotruvo. ODW. U.S. Environmental

Protection Agency Christopher DeRosa, ECAO-Cin, U.S.

Environmental Protection Agency Kurt Enslein. Genesee Computer Center, Inc..

Syracuse, NY Jeffrey Gaba. OGC, U.S. Environmental

Protection Agency Frank Gostomski, OWRS, U.S. Environmental

Protection Agency Daniel Greathouse, HERL-Cin. U.S.

Environmental Protection Agency Joseph Aroos, Tulane University, New

Orleans, LA Dev Barnes, OWRS, U.S. Environmental

Protection Agency David L Bayliss, CAG. U.S. Environmental

Protection Agency Chao W. Chen. CAG. U.S. Environmental

Protection Agency Herbert Cornish. University ofMichigan. Ann

Arbor,MI Kenneth Crump, Science Research Systems,

Inc., Ruston, LA Michael Dourson, ECAO-Cin. U.S.

Environmental Protection Agency John R. Fowle IIL, CAG. U.S. Environmental

Protection Agency R. John Garner, HERL-Cin, U.S.

Environmental Protection Agency Terrence Grady, ECAO-Cin, U.S.

Environmental Protection Agency Bernard Haberman, CAG. U.S. Environmental

Protection Agency Paul Hammond. University ofCincinnatL

Cincinnati. OH Charalingayya Hiremath. CAG, U.S.

Environmental Protection Agency Dinko Kello, Institute for Medical Research

and Occupational Health, Zagreb, Yugo. Chang S. Lao, CAG. U.S. Environmental

Protection Agency Marvin Legator, University of Texas Medical

Branch. Galveston. TX Jack McGinnity. OTS, U.S. Environmental

Protection Agency Debdas Mukerjee, ECAO-Cin. U.S.

Environmental Protection Agency Ellen O'Flaherty, University ofCincinnati.

Cincinnati, OH Fred Passman, Energy Resources Co.. Inc.

Cambridge, MA JohnF. Risher, ECAO-Cin. U.S.

Environmental Protection Agency Jeffery Rosenblatt, OTS,U.S. Environmental

Protection Agency Samuel Shibko, US FDA. Washington. DC Harry T. Skalsky, Reynolds Metal Company Todd W. Thorslund CAG, U.S.

Environmental Protection Agency Benjamin VanDuuren N.Y. University

Medical School, New York. NY James R. Withey, National Health and

Welfare Canada. Ottawa. Canada RollHartung. University ofMichigan. Ann

Arbor, MI RudolfJaeger, Harvard School of Public

Health. Cambridge, MA Arnold M. Kuzmack. CAG, U.S.

Environmental Protection Agency Si Duk Lee, ECAO-Cin, U.S. Environmental

Protection Agency LelandMcCabe, HERL-Cin, U.S.

Environmental Protection Agency Myron Mehiman, Mobil Oil Corporation, New

York, NY Jean Munson. University ofCincinnati, Cincinnati. OH Nancy Othmer, OGC.U.S. Enironmental

Protection Agency David J.Reisman. ECAO-Cin, U S.

Environmental Protection Agency Paula K. Roberson. U.S Environmental

Protection Agency L Daniel Roth. Roth Associates Dharm V. Singh, CAG, U.S. Environmental

Protection Agency Robert G.Tardiff, National Academy of

Sciences. Washington, DC Anne W. Trontell. Energy Resources Co., Inc.,

Cambridge, MA John Van Ryzin, Columbia University. New

York, NY Ronald E. Wyzga, Electric Power Research

Institute Appendix D-Response to Comments on Guidelines forDeriving Water Quality Criteria for the Protection ofAquatic Life and Its Uses
Introduction

Two versions of the Guidelines were published in the Federal Register for comment. The first version (43FR 21506, May 18, 1978 and 43 FR 29028, July 5, 1979) was simply published for

comment. The second (44 FR 15926, March 15,1979) was published as part of the request for comments on the water quality criteria for 27 of the 65 pollutants. The second version was meant to be clearer and more detailed than the first, but very similar technically. Since thetwo versions were
so similar, comments on both will be dealt with simultaneously.

Many comments were received that

no draft water quality criteria forany of the 65 pollutants should have been issued for public comment until the
comments on the first version of the Guidelines had been dealt with adequately and the Guidelines changed appropriately. The comments on the first version were read and the Guidelines

were revised in an attempt to make the second version dearer and more detailed than the first.However, an extensive revision ofthe technical
content ofthe Guidelines was not attempted between the first and second versions because the Agency was preparing water quality criteria based

on the Guidelines. The Agency could have avoided this criticism simplyby
not publishing any version ofthe Guidelines for comment until March 15. 1979, butthis would have greatly reduced the lengthof time available for people to consider the Guidelines and
comment on them.As it was, some people commented that the comment period announced on March 15,1979,

was too short.

1. Comment-The procedures used to derive criteria in the "RedBook" were upheld in court and probably should still be used.

Response-The procedures used in the Guidelines are similar to some ofthe procedures used to develop criteria in the "Green Book", "Blue Book", and "Red Book". The Guidelines are designed to be more objective and systematic,to deal more adequately with residues, and to incorporate the concept ofa minimum data base.

2. Comment-Criteria should be compilations ofcritically reviewed data with no synthesis or interpretation.

Response-Neither P.L. 92-500 nor the Consent Decree specify theform which

a criterion must take. The Consent Decree (para.11, p. 14) specifies that such criteria "shall state, inter al'a, recommended maximum permissible concentrations". Adequate precedents have been set in the "Green Book', "Blue Book', and "Red Book"for the form ofcriteria used in the Guidelines.

3. Comment-The Guidelines and" criteria should be developed by a
consensus ofaquatictoxicologists rather than by EPApersonnel only.

Response-EPA certainly wants the Guidelines and the criteria to be as good

aspossible and as acceptable to as many Interested people as possible.To this end,EPA has widely distributed draft versions ofthe Guidelines and the criteria documents, discussed themwith many people, considered the comments received, and made many significant technical changes and editorial revisions. Itis questionable whether or not a true consensus could havebeen reached by anymeans withinthe time available. In addition. EPA has a legislative responsibilitywhich it should not delegate to someone else.

4. Comment-The Guidelines should be updated regularly.

Response-The Guidelines are not being promulgated as a regulation or directive.The purpose ofpresenting these Guidelines is to show how the water quality criteria for aquaticlife

were derived for the 65 pollutants. if EPA uses these Guidelines again, they will be revised totake into account new data, concepts, and ideas.

5. Comment-The objectives, purpose, and limitations ofthe Guidelines should be stated.

Response-The introductory portion ofthe Guidelines hasbeen expanded to address these subjects more fully.

6. Comment-The Guidelines are too ambiguous.

Response-The Guidelines have been revised and rewritten, partly to improve clarity andprovide additional details. It is notpossible toprovide explicit details

on allitems; in some areas onlygeneral guidance canbe provided atthistime. 

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