### Federal Register 

# I Vol. 45, 

### No. 231 

# I Friday, November 28, 1980 

# I Notices '79347



### Residue Value**.1.ax~** 

### per.is.ible tissue concentration) (mean norBalized BCF)(appropriate pereent lipids)

a. For an FDA action level for fish oil, the appropriatepercentlipids value is 100.

b. For an FDA actionlevel for fish. the appropriate percentlipids value is 15 for freshwater criteria and 16 for saltwater criteria because FDA actionlevels are applied on a species-by-species basis to commonly consumed species. The edible portion ofthe freshwater lake trout averages about 15 percentlipids, and the edible portion ofthe saltwater Atlanticherring averages about 16 percent lipids (Sidwell. V. D.• etal. 1974 Composition ofthe Edible Portion of Raw (Fresh orFrozen) Crustaceans, Finfish. andMollusks. I. 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 or group ofaquatic species which constitute a major portion ofthe diet ofthe wildlife species.

F. The Final Residue Value is obtained by selecting the lowest ofthe available residue values. It should be noted that in many cases the Final Residue Value will not be lowenough. For example. a residue value calculated from anFDA actionlevel wouldresultin

an average concentration in the edible portion ofa fatty species that is atthe action level. On the average halfofthe individuals ofthe species wouldhave concentrations above the FDA action level. Also. the results ofmany chronic. feeding studies are concentrations that

cause adverse effects.

X. 

### OtherData

Pertinentinformation thatcouldnot beusedin earlier sections maybe available.concerning adverse effects on aquatic organisms and their uses. The
most important.ofthese are data on flavor impairment. reduction in survival, growth. orreproduction. or any other adverse effect thathas been shown to bebiologically significant. Especially important are data for species for which exceeded onthe average in a 24-hour period and one that should not be exceeded at any time during the 24·hour period. This two-number criterion is intended to identify waterquality 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 wouldhave to be to provide the same degree ofprotection.

B. The maximum concentrationis the Final Acute Value or is obtained from the Final Acute Equation.

C. The 24-hour average concentration is obtained from the Final Chronic Value. the Final PlantValue. andthe Final Residue Value by selecting the lowest available value. unless other data (see Section X) from testsinwhich the toxicant concentrations were measured show that a lower 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 PlantValue. andthe FinalResidue Value by selectingthe one thatresults in the lowest concentrations inthe normal
range ofthe waterquality characteristic. unless otherdata (see Section X) from
tests inwhich the toxicant concentrations were measured show that a lower value shouldbe used.

D. The criterionis (the 24-hour
average concentration) as a 24-hour
average andthe concentration should
notexceed (the maximum concentration) at any time.



### XII Review

A. Onthe basis ofall available pertinent laboratory and field information. determine ifthe criterion is consistent with sound scientific evidence. 

### Ifit is not. another criterion, eitherhigher or lower. should be derived using appropriate modifications ofthe Guidelines.

These Guidelines were written by Charles E. Stephan. Donald I. Mount, David J. Hansen, John H. Gentile. Gary A. Chapman and William A. Brungs of Appendix C-Guidelines and

. Methodology Used in the Preparation of HealthEffect Assessment Chapters of the ConsentDec::ree Water Criteria Documents



### I Objective

The objective ofthe health effect
assessment chapters ofthe ambient
water criteria documents is to estimate ambient waterconcentrations which do
notrepresent a significant risk to the public. These assessments should constitute a review ofall relevant information on individual chemicals or chemical classes in order to derive criteria that represent, in the case of suspect orproven carcinogens, various levels ofincremental cancerrisk. or. in the case ofotherpollutants, estimates of no-effect levels.

Ideally, ambientwater quality criteria shouldrepresent levels for compounds in ambient waterthat do not pose a hazard to thehuman population. However, in any realistic assessmentof human healthhazard. a fundamental distinction must be made between absolute safetyand the recognition of

some risk. Criteria for absolute safety wouldhave to be based on detailed

., knowledge ofdose-response .j relationships in humans. including all

sources ofchemical exposure, the types oftoxic effects elicited, the existence of thresholds for the toxic effects. the Significance oftoxicant interactions. and the variances of sensitivities and
exposure levels within the human population. Inpractice. such absolute criteria cannotbe established because ofdeficiencies in both the available data and the means ofinterpreting this information. Consequently. the individual human health effects chapters
propose criteria which minimize or specify the potential risk ofadverse human effects due to substances in ambient water. Potential social or economic costs and benefits are not 

# I

considered in the formulation ofthe criteria. 

# I

. II Types ofCriteria 

### . i
Ambient water quality criteria are 

## 1
based on three types ofbiological endpoints: carcinogenicity, toxicity (i.e., all adverse effects other than cancer]. and organoleptic effects.

For the purpose ofderiving ambient
waterquality criteria, carcinogenicity is
no otherdata 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 notbe the U.S.E.P.A. Environmental Research Laboratoriesin Corvallis. Oregon. Duluth. Minnesota. GulfBreeze. Florida. and Narragansett. Rhode Island. Numerous otherpeople. many ofwhom donotwork forU.S.E.P.A.. provided assistance and suggestions. regarded as a non-threshold phenomenon. Using this assumption, "safe" or "no effect" levels for carcinogens cannotbe established because even extremely small doses
must be assumed to elicit a finite increase in the incidence ofthe
response. Consequently. waterquality

# -

Federal Register 

## I VoL 45. No. 231 

## I Friday, November 28, 1980 

## I Notices 79348

criteria for carcinogens arepresented as

a range ofpollutant concentrations associated with corresponding incremental risks.

Forcompounds which do not manifest any apparent carcinogenic effect, the threshold assumption is used in deriving

a criterion. This assumption is basedon the premise that a physiological reserve capacity exists within the organism which is thought to be depletedbefore clinical disease ensues. Alternatively. it may be assumed that the rate ofdamage will be insignificant overthe life spanof the organism. Thus. ambient water quality criteria are derived for non­ carcinogenic chemicals. andpresumably result in no observable-adverse-,effect levels (NOAELs) in the exposed human population.



### In some instances. criteria arebased

on organoleptic characteristics, i.e., thresholds for taste or odor. Such criteria are established when insufficient information is available on toxicologic effects or when the estimate ofthe level ofthe pollutantin ambient
waterbased on organoleptic effects is lower than the level calculatedfrom toxicologic data. It shouldbe recognized that criteriabased solely on organoleptic effects do not necessarily represent approximations ofacceptable risk levels for human health.

Several ambient waterquality criteria documents dealwith classes of compounds which include chemicals exhibiting varying degrees ofstructural similarity. Bel(ause prediction of biological effects based solely on

, structuralparameters is difficult, the

derivation ofcompound-specific criteria is preferable 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 significantrisk to the public. For some chemical classes. however. a compound-specific criterion
cannot be derived for eachmember ofa class. In such instances. itis sometimes justifiable to derive a class criterion in which available data on one member of

a class may be used to estimate criteria for other chemicals ofthe-class because

a sufficientdatabase is not available for those compounds.

Forsome chemicals and chemical classes. the data base wasjudged to be insufficient for thederivation ofa criterion. 

### In those cases. deficiencies in the available information are detailed.

III. Approach

The human health effects chapters" attempt to summarize all information on the individualchemicals orclasses of chemicals which might beuseful 

### inthe risk assessmentprocess to develop waterquality criteria. Although primary emphasis is placed onidentifying 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 knownrelevant

, information. Review articles andreports

are often used inthe process ofdata evaluation and synthesis. Scientific judgment is exercisedin the review and evaluation ofthe datain each document andin 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 consistentwith respect to the format andthe application ofacceptable scieHtific principles. Evaluation procedures used in the hazard
assessmentprocess follow the principles outlinedby the National Academy of Sciences in Drinking WaterandHealth (1977) andthe 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 datawas obtainedfrom monitoring studies ofair. water. food. soil. andhuman or animal tissue residues. Themajor purpose of this sectionis 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 canbe valuable in developing and assessing a waterquality criterion. 

### In these documents exposure from consumption ofcontaminated water and contaminatedfish and shellfish products is used in criterionformulation. Data for all modes ofexposure areuseful in relating total intake to the expected contribution from contaminated water. fish, and shellfish. 

### In a.ddition. ' information for all routes ofexposure. noflimited to drinking water and fish and shellfish ingestion, canbe used to iustify or assess the feasibility ofthe formulation ofcriteria for ambient water.

The use offish consumption as an exposure factor requires the quantitation ofpollutantresidues in the edible portions ofthe ingested species. Accordingly, bioconcentration factors {BCFs} are used to relate pollutant residues in aquatic organisms to the pollutant concentrationin the ambient
watersinwhich they reside.

To estimate the average per capita intake ofa pollutant due to consumption ofcontaminated fish and shellfish the results ofa diet survey were analyzedto calculate the average consumption of freshwater and estuarine fish and shellfish'(U.S. EPA, 198O). A species is considered to be a consumed freshwater

or estuarine fish and shellfish species if
at some stage in its life cycle. itis harvested from fresh or estuarine water for human consumption in significant quantities (Stephan. 1980).

Three different procedures are used to estimate the weighted average BCF depending upon the lipid solubility of the ch,emical and the availability of bioconcentration data.

For lipid-soluble compounds. the average BCF is calculatedfrom the weighted average percentlipids inthe edible portions ofconsumedfreshwater and estuarine fish and shellfish which

was calculatedfrom dataon consumption ofeach species andits corresponding percentlipids to be 3.0 percent (Stephan, 1980).Because the steady-state BCFs for lipid-soluble compounds are proportional to percent lipids, bioconcentrationfactors for fish and shellfish canbe adjusted to the
average percentlipids for aquatic organisms consumedbyAmericans. For manylipid-soluble pollutants. there exists at leastone BCF for which the percentlipid value wasmeasuredfor the tissues for which the BCF is determined.

With 3.0 percent as the weighted average percentlipids for freshwater and estuarinefish 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 lipids for

average diet=3.0 percent Measured BCF of 17for

trichloroethylene with bluegills at 4.8 percentlipids Weighted average BCF for average

diet equals



## 17 X 

## 3.0% 

# = 10.6 

## 4.8%79349

n



### o

o

r



### le 

### 1. 

### Federal Register 

## },.yo1. 45. 

### No. 231 

## I Friday, November 28, 1980 

## I Notices

As an estimate. 10.6 is used for the

BCF. In those cases where an appropriate bioconcentration factor is not available. the equation "Log BCF=(0.85 Log P)­ 0.70" canbe used (Veith. et ai. 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 

# I 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, andin 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 otherroutes ofexposure. C. Effects

This section summarizes information

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

The major goal ofthis section is to survey the suitability ofthe data for use 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

assesspotential 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 in generating water quality criteria. Theseproblems. as . summarized by the National Academy ofSciences (NAS. 1977) are as folloWs:

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

2. Ifexposure has been ubiquitous, it
may beimpossible to assess the effects ofa 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 are usually unsatisfactory.

3. Itis usually difficult to determine doses in human exposures.

4. Usually. it is hard to identify small ' changes in common effects. which may nonetheless be important ifthe population is large.

5. Interactions in a "nature-designed" experiment usuaHy cannotbe controlled.

,

Although these problems often prevent the use ofepidemiological data inquantitative 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 conftrmed data base are considered to be reliable.

The decision to establish a criterion based on a noncthreshold model is made after evaluating all available information on carcinogenicity and supportive information onmutagenicity. The approach and conditions for the qualitative decision ofcarcinogenicity

are outlined in the U.S. EPA Interim Cancer Guidelines (41 FR 214.02). in a report by Albert. et a1. (1977). and in the Interagency Regulatory Liaison Group (IRLG) guidelines on carcinogenicrisks (IRLG. 1979). Itis assumed that a substance which induces a statistically significant carcinogenic response in animals has the capacity to cause

cancer in humans. A chemical which has not induced a significant cancer response inhumans or experimental animals is not identified as a carcinogen, even though its metabolites

orclose structural analogues might induce a carcinogenic response or itwas shown to be mutagenic in anin vitro system.

Itis recognized that some potential human carcinogens may not be identified by the guidelines given above. For example. compounds for which there is plausible but weak qualitative evidence ofcarcinogenicity in experimental animal systems (such as data from mouse skin painting or strain A mouse pulmonary adenoma) would be included in this category. The derivation ofa criterion for human consumption from these studies in not valid. regardless ofthe qualitative outcome. In addition. thereare certain compounds (e.g., nickel and beryllium] which were

. shown to be carcinogenic in humans

after inhalation exposure by chemical form. but have induced thus far no response in animals orhumans via ingesting their soluble salts. Nevertheless, a non-threshold criterion is developed for beryllium because
tumors have beenproduced in animals
at a site removed from the site of administration; in contrast. a threshold criterion is recommended for nickel because there is no evidence oftumors at sites distantresulting from administration ofnickel solutions by either ingestion orinjection.

For those compounds which were not reported to induce carcinogenic effects

or for those compounds for which carcinogenic data are lacking or insufficient. an attempt is made to estimate a no-effect level. In many respects. the hazard evaluation from these studies is similar to that of bioassays for carcinogenicity. In order
to more closely approximate conditions ofhuman exposure, preference is given
to chronic studies involving oral
exposures in water or diet over a significantportion 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 the biological endpoints used to define a no-effect level. They may 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 otherchronic effects

occur at exposure levelsbelow those causing teratogenicity. reproductive impairment. orbehavioral effects. the former are used in deriving the criterion. For most ofthe compounds evaluated thus far. teratogenicity andreproductive impairment occur at doses near

. maximum tolerated levels with dose

administration schedules well above estimated environmental exposure levels. Moreover. information on behavioral effects. which could be of

### 79350 

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

significance, is not available for most 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 10percent ofthe life span of the test organism. e.g.• 90 days for the rat with an average life
span of 30 months. Such studies 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 ofrelention 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 maybe estimated from dermal or inhalation studies. Such estimates involve approximationlfofthe total dose administered based on assumptions..

. aboutbreathing rates and/or magnitude· ofabsorption.

D. Criterion Rationale

This sectionreviews existing standards for the chemical(s), summarizes data on ·current levels of human exposure, attempts to identify special groups at risk, and defines the basisfor the recommended criterion.

Information on existing standards is includedprimarily for comparison with the proposed water quality criteria. Some ofthe present standards, such as those recommended by the Occupational Safety and Health Administration (OSHA) or the American Conference ofGovernmental Industrial Hygienists (ACGlH), 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 Drinking Waterand Health (1977) or in the U.S. EPA Interim Primary Drinking WaterStandards. are

more closely related to proposed water quality criteria. Emphasis is placed on detailing the basis for the existing standards whereverpossible. Summariesofcurrent levels ofhuman exposure. presented in this section, specifically address the suitability ofthe data to derive water quality criteria. The identification of special groups at risk, eitherbecause 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 crileria section summarizes and qualifies all ofthe data used in developing the criteria.

IV. Guidelinesfor Criteria Derivation

The derivation of water quality criteria from laboratory animal toxicity data is essentially atwo-step procedure. First, a total daily intake for humans
must be estimated which establishes either a defined level ofrisk for nOnthreshold effects or a no-effect level for threshold effects. Secondly, assumptions
must be made 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 beenmade that

a compound has the potential for causing cancers in humans and that data exist which permit the derivation ofa criterion. the water concentration which is estimated to cause a lifetime carcinogenic risk of 10- 5is determined. The lifetime carCinogenicity risk is the probability that a person wouldget

cancer sometime in his or her life assuming continuous exposure to the compound. The water concentrationis calculated by USing the low-dose extrapolation procedure proposed by Crump (1980). This procedure is an improvement on the multistage low dose extrapolation procedure by Crump. et al. (1977).

The data used for quantitative estimates are oftwo types: (1) lifetime animal studies. and (2) human studies where excess cancer risk has been associated with exposure to the agent. In animal studies itis assumed. unless evidence exists to the contrary, that ifa carcinogenicresponse 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 no really solid scientific basis for anymathematical 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 whichrisk model to use. At the present time. the dominant view ofthe carcinogenic process involves the concept that most agents which cause

cancer also cause irreversible damage to DNAThis 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 that is characteristic 

### of mutagenesis is associated with a linear non-threshold dose-response relationship. Indeed. there is substantial evidence from mutagenesis 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 inducedby aflatoxin in the diet). There

. is also some evidence from animal

experiments that is consistent with the linear non-threshold hypothesis (e.g.• liver tumors induced in mice by 2­ acetylaminofluorene in the large scale EDoI study at the National Center of Toxicological Research. and the initiation stage ofthe two-stage carcinogenesis model in the rat liver and the mouse skin).

Because it has the best, albeit limited. scientific basis ofany ofthe current ma!hematical extrapolationmodels, the linear non-threshold model has been adopted as the primary basis for risk extrapolation to low levels ofthe doseresponse relationship. The risk
assessments made with this model should be regarded as conservative, representing the most plausible upper limit for the risk; i.e., the true riskisnot likely to behigher than the estimate. but it could be smaller. 

# -

T
to ( dos is 

### tJ de.
em, be: inc inc the 95] ext wi1 ex) cal**~** Ex Ca


### M(

I (pI mt 

### p«(

wi:

Eq

wI

is 

### d(

UI 

## A

### 79351 Federal Register! VoL 45, 

### No. 231 

## I Friday, November 28, 1980 

## I Notices

The mathematical fDrmulation chDsen to' describe the linear. nDn-threshDld dDse-response relationship atlow doses is the imprDved multistage mDdel develDped by Crump (1980]. This model employs enDugh arbitrary constants to' be able to' fit almDst any monotDnically increasing dose-response data and it incorporates a prDcedure for estimating the largest pDssible linear slope (in the 

### 95 percent cDnfidence limit sense] atlow extrapDlated dDses that is cDnsistent with the data at all dDse levels ofthe experiment. Fur this reason. it may be called a "linearized" multistage model.

2. PrDcedure ofLDw-DDse ExtrapDlatiDn Based Dn Animal Carcinogenicity Data.

A. Description ofthe Extrapolation Model

Let P(d] represent the lifetime risk (prDbability} ofcancer atdose d

..The multistage model has the form p(d)=l-exp [-(q.+qtd+q.d2+ ..• +qkdk)) where:**ql~O,** andi=O, 1, 2•••••k Equivalently.

A(d)=l-exp (-(q,d+q.d2+ ..• +qkdk))
where:



# . A(d) == 

## P(d)

- P{o), 1 	- P(o)

is the extra risk overbackgroundrate at dose d.

The point estimate ofthe coefficients ql. i=O. I, 2•..• 

### ,k. and consequently the extrarisk function A(d) at anygiven dDse d. is calculatedby maximizing the likelihoodfunctiDn ofthe data.

The point estimate andthe 95percent upper confidence limit ofthe extra risk A(d) are calculatedbyusing the computer program GLOBAL 79 developedby Crump andWatson (1979). Upper 95percent confidence limits on

> the extrarisk andlower 95 percent cDnfidence limits on the dDse producing a given risk are determined frDm a 95 percent upper confidence limit. qt·, on parameterql. Whenever ql;cO. at IDW dDses extra riskA(d) has apprDximately the form A(d)=qlxd.Therefore. qlXa is a 95percentupperconfidence limit on the extra risk and R/ql

### * is a 95 percent lowerconfidence limit on the dose producing an extra risk of

### R. Let 

### Lo be the maximum value ofthe 10g-likelihoDd function. Theupperlimit ql* is calculatedbuyincreasing ql to a vah,le qt· such thatwhen the log-likelihood is again maximized subject to this fixed value ql

### * for the linear coefficient, the resulting maximum value ofthe IDglikelihood Ll satisfies the equatiDn 20"..0-Lt)=2.70554 where 2.70554 is the cumulative 90 percentpoint ofthe chi-square distribution with one degree offreedom. which correspDnds to a 95 percentupper limit (Dne-sided). This approach of cDmputing the upper cDnfidence limit for the extra risk A(d] is an improvement on the Crump, et al. (1977} model. The
upper confidence limit for the extra risk calculated atlow dDses is always linear. This is conceptually consistent with the linearnonthreshold cDncept discussed earlier. The slope q,. is taken as an upperbDund Dfthe potency ofthe cpemical in inducing cancer at IDW doses.

In fitting the dose-response model. the number ofterms in the polynomial g is chosen equal to [h-1). where h is the number ofdose groups in the experiment. including the control group.

Whenever the multistage mDdel dDes nDt fit the data sufficiently. data at the highest dose is deleted andthe model is refitted to therestofthe data. This is continued until an acceptable fit to the data is obtained. To determine whether

ornDt a fit is acceptable. the chi-square statistic: 

### h 

## (Xi

- NiP i)2 X2 = 

## NiPi 

## (1

- Pi) i=1

. is calculated. where NI is the number of animals in the ilb dose group. XI is the number ofanimals in the jib dDse group with a tumDrresponse. PI is the probability ofa response in the ith dose
group estimatedbyJitting the multistage model to the data. andh isthenumber ofremaining groups.

The fit is determined to be • unacceptable wheneverchi-square (XI) is larger than the cumulative 99percent point ofthe chi-square distributiDn with fdegrees offreedom, where fequals the number ofdose groups minus the number ofnon-zerO' multistage cDefficients.

3. Selection andForm ofData used to Estimate Parameters in the ExtrapDlation Model.

FDr some chemicals. several studies in different animal species. strains. and

sexes each conducted at several dDses and different routes ofexposure are available. A chDice must bemade as to which ofthe data sets from several studies are to beused in the model. Itis alsO' necessary to CDrrect fDr metabolism differencesbetween species-and for differences inabsorption via different
routes ofadministratiDn. The procedures. listed below, used in evaluating these data are cDnslstent with the estimateof

### a maximum-likely

.. risk. a. The tumor incidence data are separated according to organ sites or tumor types. The set data (i.e., dose and
tumorincidence) used in the model is
set where the incidenceis statistically significantly higher than the control fDr
at leas"Dne testdose level and/or where the tumor incidence rate shows a statistically significant trend with respect to dose level. The data set which gives the highest estimate oflifetime carcinogenic risk qt· is selected in most

cases. HDwever. efforts are made to exclude data sets which produce spuriDusly high risk estimates because ofa small number ofanimals. That is. if two sets ofdata ShDW a similardDseresponserelationship and one has a very small sample size, the set ofdata which has the larger sample size is selected fDr calculating the carcinogenic potency.

b. Ifthere are two ormore data sets of comparable size which are identical with respect to species. strain. sex. and
tumor sites, the geometric mean Dfql•• estimatedfrom each ofthese data sets is used for risk assessment. The geometric

meanofnumbers Al•A,•••••Am is defined as (AlxA,x ••• XAmP/m

c. Ifsufficient data exist for two or

more significant tumor sites in the same study, the number Dfanimals with at least one ofthe specific tumor sites under consideration is used as incidence data in the model.

d. Following the suggestion ofMantel andSchneiderman {1975]. we assume that mg/surface area/dayis an equivalent dosebetween species. Since
to a close approximation the surface

area is proportional to' the o/ards power ofthe weight as wouldbe the case for a perfect sphere, the expO'sure in mgto/srds
power ofthe body weight/day is similarly considered to beanequivalent exposure. In ananimal experiment. this equivalent dDse is computed in the following manner: Let:

L.=duration ofexperiment I.=durationofexposure m=average dose perday in mg during adrilinistration ofthe agent (I.e.• during 

### 1.) W

### =average weight ofthe experimental anilnal.
Then. the lifetime average exposure is 

## le X 

## m d •

## ----:O=~ Le X 

## W 2/3

Often exposures are**~Dt** given in units ofmg/day. and itbecomes necessary to'
convert the given exposures into mgt day. FDr example. in most feeding studies. exposure is expressed asppm in the diet. Inthis case the exposure 

### (mgl day) is derivedby: m=ppm X F X r

### 79352 Federal Register 

# rVol. 45, No. 231 

## I Friday, November 28. 1980 

## I Notices

-== where ppm is parts permillion ofthe Case 1 ofspecies, itfollows that m=k W2/sx v carcinogenic agent in the diet. F is the weight ofthe food consumed per day in kgm,s. andr is the absorption fraction.

In the absence ofany data to the contrary. r is assumed to be one. For a uniform diet the weight ofthe food consumed is proportional to the calories

:d**~** : required. which, in tum. is proportional
to the surface area or the %rds power of the weight. so that: mappmX \y2/3Xr or



### m ppm Cl 

### rW2/3

As a result. 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 to moisture content differences. Instead. we use an empirically derived food factor, f=F/W, which is the fraction ofa species body weight that is consumed per day as food. We use the rates given below.

species w

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 mg/W21 3is



## m 

# =ppm x F 

# =

, Ii " 

### r 

## x 

## W 2/3 W2/3



# .. ppm x fxW 

# =ppm x f x Wl/3

. W2/3
When exposure is given in terms of mg/kg/day=m/Wr=s the conversionis simply:



## m

.. S X 

## Wl/3 rW2/3

When exposure is via inhalation. the calculation ofdose can be 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 ofairbreathed in, and (2) where the carcinogen is a poorly water-soluble gas which reaches an equilibrium between the airbreathed and the body compartments. After equilibrium is reached. the rate ofabsorption ofthese agents is expected to be proportional to metabolic rate, which in tum is proportional to therate ofoxygen consumption, which in turn is a function ofsurface area. Agents that are in the form of particulate matter orvirtually completely absorbed gases such as SO,

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=IXvXrwhere I is inhalation rate per day in mS. v is mg/m3ofthe agent in air, and r is the absorption fraction.

The inhalation rates. I, for various species can be 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 ofother weights. W. (expressed in kg). the surface area proportionality can be used to determine breathing rates (in

m3 /day) as follows:

. For mice. 1=0.0345 (W/0.025)21 

### 'm3 / day

For rats. 1=0.105 (W/0.113)2/

### 'ms 

### /day For humans, the values of20 m3 /day· is adopted as a standard breathing rate (ICRP. 1977).

The equivalent exposure in mg/W2/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:

i=1I Species 

### w W

Man

............................................................................ 70 0.29 Rat

............................................................................. 0.35 0.64 Mice

...............................................,............._......._. 0.03 1.3

Therefore. for particulates or completely absorbed gases. the equivalent exposure in mg/W2T3 is:



### m • Ivr. iWvr • iwl/3 vr W2/3 W2/3 W2/3

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

Case 2

The dose in

## mg/day ofpartially soluble vapors is proportional to the**~** consumption which in turn is proportional to W2/sand to the solubility ofgas in body fluids, which

can be expressed as an absorption coefficient r for the gas. Therefore, when expressing the 0, consumption as O,=k W2/3, where k is a constant independent

• From "Recommendation ofthe International Commission onRadiological Protection." page 9. the average breathing rate is 10'em'per &-hour work day and ZX10'em'in Z4 hours. x r or


### d.. m

.. kvr 

## W2/3'

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./m3in 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 similarin

man and animals (Dripps. et a1. 1977). When the animals were exposed via the oral route and human exposure is via inhalation or vice-versa, the assumption is made. unless there is pharmacokinetic evidence to the contrary, that absorption is equal by either exposure route.

e. Ifthe duration ofexperiment 

### (1...) is less than the natural·life span ofthe test animal (L). the slope ql·, or more generally the exponent g(d). is increased by multiplying a factor 

### (L/L.P. We

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

cancer will continue to increase as a constant function ofthe background
rate. The age specific rates for humans increase at least by the 2nd power ofthe
age and often by a considerably higher power. as demonstrated by Doll (1971). Thus, we would expect the cumulative
tumor rate to increase by at least the 3rd power ofage. Using this fact. we assume that the slope ql

### *, or more generally, the exponent g(d), would also increase by at least the 3rdpower ofagEf. As a result, if the slope ql· [or g(d)) is calculated at age 

### 1.... we would expect that ifthe experimenthadbeen continued for the full life span. L. at the given average exposure. the slope ql* [or g(d)) would have been increased by at least 

### (L/L.P.

This adjustment is conceptually consistent to the proportional hazard model proposed by Cox (1972) and the tim!!-to-tumor model considered by Crump, et al, (1977) where the probability ofcancer at age t and dose d is given by P(d.t)=l-exp[-f(t)xg(d)]

4. Calculation ofCarcinogenicPotency Based on HumanData. 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 ofcancerrates on lifetime average dose. which is equivalent to the factor ql•. 

### Ifthey show

### 79353 Federal Register IVoI. 45. 

### No.- 231 

## I Friday. November 28. 1980 

## I Notices



### v

e

In

S



### e

1

IC in



### t

d

e

d

.e

e It if



### y

.e 

### l

'i no carcinogenic effect whenpositive animal evidence is available. then itis assumed that a risk does existbutit is smaller than could have been observed in the epidemiologic study, and anupper limit ofthe cancer incidence is calculated assuming hypothetically that the true incidence is justbelow the level ofdetection in the cohort studied. which is determined largely bythe cohort size. Wheneverpossible. human data are used in perference to animal bioassay data.

Inhuman studies, the response is measured in terms ofthe relative risk of the exposed cohort ofindividuals compared to the control group. Inthe 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. andthat it is the same for all ages. Itfollows that the carcinogenicpotency is equal to

_ [R(X)-l]/Xmultiplied bythe lifetime risk at that site in the generar population. Except for anunusually well-documented human study, the confidence limit for the excess risk is
not calculated. due to the difficulty in accounting for the uncertainty inherent inthe data (exposure and cancer response).

5. Calculation ofWater Quality Criteria. After the value ofql" in (mg/ kg/day)"'1 has been determined. the lifetime risk. P. from an average daily
exposure ofx mg/kg/day is found from the equation P=ql·x. Therefore. ifthe lifetime risk is set atP=10- 5for calculation purposes. the intake. I, in mg/day for a 70 kg person canbefound bythe equation: 1=70X10- 5/ql" The intake ofthe agent from ambient
water is assumed to come from two

sources: (1) drinking an average of2 liters ofwaterperday, and (2) ingesting

an average of6.5 grams offish perday. 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)XR(l/kg fish)X0.0065 kg fish/day+ C(mg/lX21/ day=C(2+0.0065R) where Cis the
water concentrationin mg/1. Therefore. the waterconcentration inmg/l

. corresponding to a lifetime risk of10-a for a 70 kg person is calculated bythe formula: 

## 70 X 

## 10-5 

# c

.. 01*(2 + 

# 0.0065 R} B. ThresholdEffects
1. Use ofAnimal Toxicity Data (Oral). In developing guidelines for deriving criteriabased on noncarcinogenic
responses, five types ofresponse levels

are considered:

NOEL-No-Observed-Effect-Level NOAEL-No-Observed-Adverse-Effect-Level LOEL-Lowest-Observed-Effect-LevelWAEL-Lowest-Observed-Adverse-EffectLevel· FEL-Frank-Effect-Level

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

One ofthe major problems encountered in consideration ofthese conceptsregards the reporting of "observed effect levels" as contrasted to "observedadverse effect levels". The
terms "adverse" vs. "notadverse" 

### are at times satisfactorily defined. but due to increasingly sophisticated testing protocols. more subtle responses are being identified. resulting in a need for judgmentregarding 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. anda clearly defined FEL in relatively closely spaced doses•.the threshold region has beenrelatively well defined; such data

are very useful for the purpose of deriving a criterion. On the otherhand. a clearly defined FEL has little utility in establishing criteria when itstands alone, because such a level gives no indication how far removed 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 NOEL may be many orders ofmagnitude below the threshold region.

Based on the above dose-response classU'ication system. the following guidelines for deriving criteria have been adopted:

a. A free-standiilg FEL is unsuitable for the deriva:tion ofcriteria.

b. A free-standing NOEL is unsditable for the derivation ofcriteria. 

### Ifmultiple NOELs are available without additional data on LOELs, NOAELs. orLOAELs.

. the highestNOEL should be used to derive a criterion.

c. A NOAEL. LOEL, 01'LOAEL canbe suitablefor criteria derivation. A well· defined NOAEL from a chronic (at least OO-day) studymaybe used directly, applying the appropriate uncertainty factor. For a LOEL. ajudgment needs to be made whether it actually corresponds
to a NOAEL or a LOAEL. In the case of

a LOAEL. an additional uncertainty factor is applied; the magnitude ofthe additional uncertaintyfactor is judgmental and sho\lld lie in the range of 1 to 10. Caution mustbe exercised not to substitute "Frank-Effect-Levels" for"Lowest-Observable-Adverse-EffectLevels". '

d. 

### Iffor reasonably closely spaced doses only a NOEL and a LOAEL of equal quality are available. then the appropriate uncertainty factor is applied
to the NOEL.

In using this approach, the selection andjustification ofuncertainty factors

are critical. The basic definition and guidelines for using uncertainty factors has beengiven bythe National Academy ofSciences (1977). "Safety Factor" or "Uncertainty Factor" is defined as a number that reflects the degree oramount ofuncertainty that
must be considered when experimental data in animals are extrapolated to man. When the quality and quantity of experimental data are satisfactory. a lowuncertainty factor is used; when data is judged to be inadequate or equivocal. a larger uncertainty factor is used. The following general guidelines havebeen adopted in establishing the uncertainty factors:

a. Valid experimental results from**s~udies** onprolonged ingestion by man. with no indication ofcarcinogenicity. UncertaintyFactor=10

b. Experimentalresults ofstudies of human ingestion not**availabl~** or scanty (e.g.• acute exposure only) with valid results oflong-term feeding studies on experimental animals. orin the absence ofhuman studies. valid animal studies

on one ormore species. No indication of carcinogenicity. UncertaintyFactor= 100

c. No long-term or acute human data. Scanty results on experimental animals with no indication of carcinogenicity. Uncertainty Factor= 1.000 Considerablejudgment must be used in selecting the appropriate safetyfactors 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 factor is used. Such anintermediate uncertaintyfactor maybe developed based on a logarithmic scale (e.g

.. 33. being halfwaybetween 10 and 100 on a logarithmic scale).

In determining the appropriateuse of the uncertainty factors, the phrase "no

### Federal Register 

## I Vol. 45. No. 231 

## I Friday.- November 28. 1980 

## I Notices· 79354

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 are not used either for derivation of numerical 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 for both the selection ofthe safety factor and the data to which it is applied. This discussion should reflect a critical review of the available data base. Factors to be considered include number ofanimals. species. and parameters tested; quality of controls; 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). For carcinogens. 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 criterion (C)

. is: G=ADI-(DT+IN)/[21+(0.OO65 kg

x R)] where 21 is assumed daily water consumption. 0.0065 kg is assumed daily fish consumption. R is bioconcentration factor in units ofI/kg. DT is estimated non-fish dietary intake. and IN is estimated daily intake by inhalation.

Ifestimates ofIN and DT cannot be provided from experimental data. an assumption must be made concerning total exposure. It is recognized that either the inability to estimate DT and IN due to lack 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 of scientific validity. the accurate estimate of the Acceptable Daily Intake is the major factor in satisfactory derivation ofwater quality criteria.

3. Use ofThreshold Limit Values or Animal Inhalation Studies. Threshold Limit Values (TLVs) are established by the American Conference of Governmental and Industrial Hygienists (ACGIH) and represent a.hour timeweighted 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. each TLV must be carefully examined to determine ifthe basis of the TLV contains data which can be used directly to derive a water 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
to protect 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 the basis ofthe criterion. consideration ofthe above. points is explicitly stated in the criterion derivation section of the document. Particular emphasis is placed on the quality ofthe TLV relative to the available toxicity data that normally is given priority overTLVs or similar established values. Ifthe TLV can be justified as the basis for the cirterion. then the problems associated with the estimation ofacceptable oral doses from inhalation data must be addressed.

Estimating equivalencies ofdoseresponse relationships from oneroute 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 other routes 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 can be 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 metaboUc 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 pharmacokinetic 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 to justify pharmacokinetic modeling. Such alternative approaches. however. provide less satisfactory approximations because they are not based on pharmacokinetic data. Consequently. in using the Stokinger and Woodward or related models. the uncertainties inherent in each ofthe assumptions and the basis of each assumption must be clearly stated in the derivation of the criterion.

The use ofdata pertaining to other routes of exposure 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. Ifsimplifying assumptions are used. their bases and limitations must be clearly specified.

Because of the 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 of the available data against a perceived risk to the human population.

Ifthe Stokinger and Woodward (1958) approach is used to calculate an ADI from a TLV. the general equation is: ADI=TLVxBRxDExdxAA/(AoxSF) where: ADI=Acceptable daily intake in mg TLV=Concentration in air in mg/m3 DE=Duration ofexposure in hours per day d=5days/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=CA XDExdxAAxBRX70 kgl (BWAXAoXSF) 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 m3 70 kg=Assumed human body weight BWA=Body weight ofexperimental animals in kg

### 79355 Federal Register 

## I Vol. 45. 

### No. 231 

## I Friday; 

### November 28, 1980 

## I Notices

Ao=Efficiency of·absorption from oral exposure SF= Safety factor following guidelines given above.
More formal 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 

## rug/ surface area/dayconversion factor is
not used in the derivation oftoxicity basedcriterion.

C. Organoleptic Criteria

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 human health 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 oforganoleptic data is often significantly less than that oftoxicologic data used in establishing othercriteria. Consequently. a critical evaluation of the available organoleptic data must be made and the selection ofthe most appropriate data base for the criterion
must be based on sound scientific judgment.

Organoleptic criteria are not based on toxicologic information and have no direct relationship to potential adverse human health effects. Although sufficiently intense organol-eptic characteristics could result in depressed fluid intake which. in turn. might aggravate a variety offunctional 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 derivedsolely from organoleptic data are based upon aesthetic qualities only.

Since organoleptic andhuman health effects criteria are based on different endpoints. a distinction must be made between these two sets_ofinformation. In criteria summaries involving Itoth types ofdata. the following format is used:

For comparison purposes. two approaches Were used to derive criterion levels for

-. Based on available toxicity data. for the protection ofpublic health the derived level is --.Using available organoleptic data. for controlling undesirable taste and odor quality ofambient water the estimated level is --. It should be recognized 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 cannot be derived. the following statement is to be appropriately inserted:

Sufficient data are not available for

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

D. Criteriafor Chemical Classes

A chemical class is broadly defined as anygroup 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 of a class. It involves the use ofavailable data on

one or more chemicals ofa 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 ofa class rather than to the sum ofthe 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., differentpatterns of environmental transport and degradation) on the validity ofthe criterion. Hthese effects cannotbe, assessed. an appropriate statement of

, uncertainty should accompany the

criterion.

Since relatively minor structur.al changes within a class ofcompounds

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. scientificjudgment may suggest

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

The development of a class criterion takes into consideration the following: 1. A detailed review of the chemical and physical properties ofchemicals within the
group should be made. A close relationship within the class with respect to chemical acitivity would suggest a similarpotential to reach common biological sites within tissues. Likewise. similar lipid solubilities would suggest the possibility ofcomparable absorption and tissue distribution.

2. Qualitative and quantitative data for chemicals within the group are examined. Adequate toxicologic data on a number of compounds within a group provides a more reasonable basis for extrapolation to other chemicals of the same class than minimal data on ope chemical or a few chemicals within the group.

3. Similarities in the nature of the toxicologic response to chemicals in the class provides additional support for the prediction that the response 10 olher members of the class may be similar. 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 to other members of that class is not appropriate.

4. Additional support for the validity of extrapolation ofa criterion to other members of a class could be provided by evidence of similar metabolic andpharmacokinetic data" for Borne members ofthe class.

Based on the above considerations. 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 ofa class but no criterion for others). While no 

### a priorilimits 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 or overestimating risk/ safety and should be rigorously avoided. Although scientificjudgment has a properrole in the development ofclass criteria, such criteria are useful and defensible only ifthey are based on**adequate~ata** and scientific reasoning. The definition ofsufficient data on similarities in physical, chemical, pharmacokinetic.-or toxicologic properties to justify a class criterion may vary markedly depending on the degree of structural similarity and the gravity ofthe per:ceived risk. Consequently. it is imperative that the criterion derivation section of each document in which a class criterion is recommended explicity address each of the key issues discussed above. and define. as clearly as possible. the

### 79356 Federal Register 

## l Vol. 45, No. 231 

### I l'rlOay, 1'10VeIJIuer "'0,**~~ou** I**!~v** .....c"

limitations ofthe proposed criterionas well as the type of daJa 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 concel'li;'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 for potential 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. EssentialElements

Some chemicals. particularly certain metals, are essential to biological organisms at low levels but may be toxic andl orcarcinogenic 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 to be essential butfor which supportive data

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

To modify the toxicity and carcinogenicity based criteria, essentiality mustbe 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 ADls for non-carcinogens defines the spread of daily doses from which the criterion may be derived. Because errors

are inherent in defining both essential and maximum tolerable levels, the criterion is derived from dose levels

near the center of such a dose range. The decision to use either the MDR or RDA is guided by the spread of the doses and the quality ofthe essentiality and toxicity estimates.

The modification of criteria by consideration of essentiality must take

. into account aU routes ofexposure. 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
may be adjusted ifthe derived criterion represents a small fraction ofthe ADl or MDR. On a case-by-case basis, the modification in the use ofthe guidelines
may include the use ofdifferent safety factors for non-carcinogens or other modifications which can be explicitly justified.



### F. Use of Existing Standards

For some chemicals for'which criteria

are to be established, drinking water standards already exist. These standard's represent not only a critical

assessment ofliterature, but also a body ofhuman experience since their promulgation. Therefore, it is valid to ' accept the existing standard unless there is compelling evidence to the contrary. This decision should be made after considering the existing standards

vs. new scientific evidence whichhas accumulated since the standards have been established. There are several instances where the peer review process recommended usage ofthe present drinking water standards.**Bibliog~aphy**

Albert. R.E., et al. 1977. Rationale developed by the Environmental Protection Agency for the assessment of carcinogenic risks. Jour. Natl. Can. Inst. 58: 1537.

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

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

Crump. KS. 1980. An improved procedure for low-dose carcinogenic risk assessment from animal data. Jour. Environ. Path. Toxico!. (In press).

Crump. KS. andW.W. Watson. 1979, GLOBAL 79. A FORTRAN program to extrapolate dichotomous animal carcinogenicity data to low dose. Nat!. Inst. Environ. Health Sci. Contract No. 1-ES-2123.

Crump, KS., et al. 1977. Confidence intervals and test ofhypotheses concerning dose·response relations inferred from animal carcinogenicity data. Biometri£s. 33: 437.

Doll. R. 1971. Weibull distribution of

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

Dripps, R.D

.. et al. 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. III. P.L. Altman and D.S. Dittment (eds.] Federation ofAmerican Societies for Experimental Biology. Bethesda, Maryland. Library ofCongress No. 72-87738. Guess, H.. et al. 1977. Uncertanity 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 ofthe ICRP, Pub!. No. 26. adopted January 17. 1977. Pergamon Press, Oxford. England.

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

McNamara. B.P. 1976. Concepts in Health Evaluation ofCommercial and Industrial Chemicals. In: M.A. Mehlman, et a!. (eds.). Advances in Modern 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. 1976. Guidelines for carCinogen bioassay in small rodents. NCI Carcinogenesis Technical Report Series, No. 1. February. 1976.

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

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

Stokinger. M.E. and R.L. Woodward. 1958. Toxicologic methods for establishing drinking
water standards. Jour. Am. Water Works Assoc. 50: 517.

U.S. EPA. 1976. National Interim Primary Drinking WaterRegulations. Office ofWater Supply. EPA 570/9-76-003. U.S. Envlran. Prot. Agency.

U.S. EPA. 1980. Seafood consumption data analysis. Stanford Research Institute International. Menlo Park, California. Final report, Task 11. Contract No. 68-{)1-3687.

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

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

Health Effects Methodology-Authors, Contributors and 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

School. New York, NY

. Patrick R. Durkin. Syracuse Research Corporation, Syracuse. NY Steven D. Lutkenhoff, ECAO, U.S.

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

Protection Agency Edward Calabrese, University of

Massachusetts, Amherst, MA Thomas Clarkson. University ofRochester,

Rochester,NY

### 79357 Federal**Re~ilrter** 

### tVol. 45, No. 231'/ 

### Friday, November 28, 1980 / 

### Notices



# ! 

# I

\ Joseph Cotruvo. ODW, U.S. Environmental Protection Agency ChristopherDeRosa. ECAO-Cin. U.S.

Environmenta! Protection Agency Kurt Enslein. Genesee Computer Center. Inc

..

Syraeuse. 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 Arcos. 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. MJ KennethCrump. Science Research Systems.

Ine

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

Environmental Protection Agency John R. Fowle III; CAG. U.S. Environmental

Protection Agency R. John Garner. HERL-Cin. U.s.

Environmental Protection Agency**T~rrence** Grady, ECAO-Cin. U.S. Environmental Protection Agency Bernard Haberman. CAG. U.S. Environmental

Protection Agency Paul Hammond. University ofCincinnati.

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 ofTexas Medical

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

Protection Agency Debdas Mukerjee. ECAO·Cin. U.S.

Environmental Protection Agency Ellen O·F1aherty. University ofCincinnati.

Cincinnati. OH FredPassman. Energy Resources Co

.. Inc.

Cambridge. MA

.. John F. Risher. ECAO·Cin. U.S.

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

Prote<:tion 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 RolfHartung. University ofMichigan. Ann

Arbor. MI RudolfJaeger. HarvardSchool ofPublic

Health. Cambridge, MA Arnold M. Kuzmac.k. CAG. U.S.

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

Protection Agency Leland McCabe, HERL-Cin. U.s.

Environmental Protection Agency Myron Mehlman. Mobil Oil Corporation. New

York. NY Jean Munson. University ofCincinnati. Cincinnati. OH Nancy Othmet. OGC. U.S. Environmental

Protection Agency David

### J. Reisman. ECAO-Cin. U.S.

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

Protection Agency H. 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 for DerivingWater Quality 

### Criteria for the Protection ofAquatic Life and Its Uses

IntroducUon



### Two versions ofthe Guidelines were published in the Federal Register for
comment. The first version (43 FR 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 

### ofthe 65 

### pollutants. The second version was meant to be clearer and more detailed than the first. but very similar technically. Since the two versions were

so similar, comments on both will be dealt with simultaneously.

Many comments were received that

no draft water quality criteria for any of the 65 

### pollutants should have been issued for public comment until the
comments on the first version ofthe Guidelines hadbeen 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 dearerand 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 waterquality criteria based

on the Guidelines. The Agency could have avoided this criticism simply by
not publishing any version ofthe Guidelines for comment until March 15, 1979. 

### but this would have greatly reduced the length oftime available for people to considerthe Guidelines and
comment on them. As it was. some people commented that the comment period announced onMarch 15, 1979.



### was too short.

1. Comment-Theprocedures used to derive criteria in the "Red Book" were upheld in court and probably should still be used.

Response-The procedures used in the Guidelines are similar to some ofthe procedllres 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-NeitherP.L. 92-500 

### nor the Consent Decree specify the form which

a criterion must take. The Consent Decree (para. 11, 

### p. 14) specifies that such criteria "shall state, interalia. 

### 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 ofaquatic toxicologists rather than by EPA personnel only.

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

as possible 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 them with many people. considered the comments received. and made many significant technical changes and editorial revisions. It

### is questionable whether.or not a true consensus could have been reachedby any means within the time available. In addition, EPA has a legislative responsibility which 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 aquatic life

were derived for the 65 

### pollutants. If 

### EPA uses these Guidelines again, they willberevised to take 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 has been expanded to address these subjects more fully.

6. Comment-The Guidelines are too ambiguous.

Response-The Guidelines have been revised andrewritten. partly to improve clarity and provide additional details. It 

### is not possible to provide explicit details

on all items: in some areas only general guidance canbe provided at this time.