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

Residue Value - (axim m permissible tissue concentration)

(mean normalized

a. For an FDA action level for fish oil,

the appropriate percent lipids value is

100.

b. For an 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 of the freshwater lake trout

averages about 15 percent lipids, and

the edible portion of the saltwater

Atlantic herring averages about 16

percent lipids (Sidwell, V. D., et al. 1974

Composition of the Edible Portion of

Raw (Fresh or Frozen) 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 of an

aquatic species or group of aquatic

species which constitute a major portion

of the diet of the wildlife species.

F. The Final Residue Value is

obtained by selecting the lowest of the

available residue values. It should be

noted that in many cases the Final

Residue Value will not be low enough.

For example, a residue value calculated

from an FDA action level would result in

an average concentration in the edible

portion of a fatty species that is at the

action level. On the average half of the

individuals of the species would have

concentrations above the FDA action

level. Also, the results of many chronic

feeding studies are concentrations that

cause adverse effects.

X. Other Data

Pertinent information that could not

be used in earlier sections may be

available concerning adverse effects on

aquatic organisms and their uses. The

most important of these are data on

flavor impairment, reduction in survival,

growth, or reproduction, or any other

adverse effect that has been shown to

be biologically significant. Especially

important are data for species for which

no other data are available. Data from

behavioral. micorcosm, field, and

physiological studies may also be

available.

XI. Criterion

A. The criterion consists of two

concentrations, one that should not be

BC)(appropriate per'ent lipids)

exceeded on the 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 water quality

conditions that should protect aquatic

life and its uses from acute and chronic

adverse effects of both cumulative and

noncumulative substances without being

as restrictive as a one-number criterion

would have to be to provide the same

degree of protection.

B. The maximum concentration is the

Final Acute Value or is 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 Section X) from tests in which

the toxicant concentrations were

measured show that a lower value

should be used. If toxicity 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 in which 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 if the criterion is

consistent with sound scientific

evidence. If it is not, another criterion,

either higher or lower, should be derived

using appropriate modifications of the

Guidelines.

These Guidelines were written by

Charles E. Stephan, Donald I. Mount,

David 1. Hansen, John H. Gentile, Gary

A. Chapman and William A. Brungs of

the U.S.E.P.A. Environmental Research

Laboratories in Corvallis, Oregon,

Duluth, Minnesota, Gulf Breeze, Florida,

and Narragansett, Rhode Island.

Numerous other people, many of whom

do not work for U.S.E.P.A., provided

assistance and suggestions.

Appendix C-Guidelines and

Methodology Used in the Preparation of

Health Effect Assessment Chapters of

the Consent Decree Water Criteria

Documents

L Objective

The objective of the health effect

assessment chapters of the ambient

water criteria documents is to estimate

ambient water concentrations which do

not represent a significant risk to the

public. These assessments should

constitute a review of all relevant

information on individual chemicals or

chemical classes in order to derive

criteria that represent, in the case of

suspect or proven carcinogens, various

levels of incremental cancer risk, or, in

the case of other pollutants, estimates of

no-effect levels.

Ideally, ambient water quality criteria

should represent levels for compounds

in ambient water that do not pose 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 of dose-response relationships in humans, including all

sources of chemical exposure, the types

of toxic effects elicited, the existence of

thresholds for the toxic effects, the

significance of toxicant interactions, and

the variances of sensitivities and

exposure levels within the human

population. In practice, such absolute

criteria cannot be established because

of deficiencies in both the available data

and the means of interpreting this

information. Consequently, the

individual human health effects chapters

propose criteria which minimize or

specify the potential risk of adverse

human effects due to substances in

ambient water. Potential social or

economic costs and benefits are not

considered in the formulation of the

criteria.

1I. Types of Criteria

Ambient water quality criteria are

based on three types of biological

endpoints: carcinogenicity, toxicity (i.e.,

all adverse effects other than cancer),

and organoleptic effects.

For the purpose of deriving ambient

water quality criteria, carcinogenicity is

regarded as a non-threshold

phenomenon. Using this assumption, "safe" or "no effect" levels for

carcinogens cannot be established

because even extremely small doses

must be assumed to elicit a finite

Increase in the incidence of the

response. Consequently, water quality

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

criteria for carcinogens are presented as

a range of pollutant concentrations

associated with corresponding

incremental risks.

For compounds which do not manifest

any apparent carcinogenic effect, the

threshold assumption is used in deriving

a criterion. This assumption is based on

the premise that a physiological reserve

capacity exists within the organism which is thought to be depleted before

clinical disease ensues. Alternatively, it

may be assumed that the rate of damage will be insignificant over the life span of

the organism. Thus, ambient water

quality criteria.are derived for noncarcinogenic chemicals, and presumably

result in no observable-adverse-effect

levels (NOAELs) in the exposed human

population.

In some instances, criteria are based

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

of the level of the pollutant in ambient

water based on 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 of acceptable

risk levels for human health.

Several ambient water quality criteria

documents deal with classes of

compounds which include chemicals

exhibiting varying degrees of structural

similarity. Because prediction of

biological effects based solely on

structural parameters is difficult, the

derivation of compound-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 significant risk to the

public. For some chemical classes,

however, a compound-specific criterion

cannot be derived for each member of a

class. In'such instances, it is sometimes

justifiable to derive a class criterion in

which available data on one member of

a class maybe used to estimate criferia

for other chemicals of the class because

a sufficient data base is not available

for those compounds.

For some chemicals and chemical

classes, the data base was judged 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 might be useful in the

risk assessment process 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 of human 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 of data

evaluation and synthesis. Scientific

judgmefit is exercised in the review and

evaluation of the data in each document'

and in the identification of the adverse

effects against which protective criteria

are sought. In addition, each of these

documents is reviewed by a peer

committee of scientists familiar with the

specific compound(s). These work.

groups evaluate the quality of the

available data, the completeness of the

data summary, and the validity of the

derived criterion.

In the analysis and organization of the

data, an attempt is made to be

consistent with respect to the format

and the application of acceptable

scientific principles. Evaluation

procedures used in the hazard

assessment process follow the principles

outlined by the National Academy of

Sciences in Drinking Water andHealth

(1977) and the guidelines of the

Carcinogen Assessment Group of the

U.S. EPA.

A. Exposure

The exposure section of the health

effects chapters reviews known

information on current levels of human

exposure to the individual pollutant

from all sources. Much of the data was

obtained from monitoring studies of air,

water, food, soil, and human or animal

tissue residues. The major purpose of

this-section is to-provide background

information on the contribution of water

exposure relative to all other sources.

Consequently, the exposure section

includes subsections reviewing different

routes of exposure 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

of contaminated 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 of tie

formulation of criteria for ambient

water.

The use of fish consumption as an

exposure factor requires the

quantitation of pollutant residues in the

edible portions of the ingested species.

Accordingly, bioconcentratlon factors

(BCFs) are used to relate pollutant

residues in aquatic organisms to the

pollutant concentration in the ambient

waters in which they reside.

To estimate the average per capita

intake of a pollutant due to consumption

of contaminated fish and shellfish the

results of a 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 or estuarine water

for human 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 BCF is calculated from the

weighted average percent lipids in the

edible portions of consumed freshwater

and estuarine fish and shellfish which

was calculated from data on

consumption of each 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 for which 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 of 17 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 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" can be used (Veith, et al. 1979] to

estimate the BCF for aquatic organisms

containing about 7.6 percent lipids

(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.

For non-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 of the

average diet.

B. Pharmacokinetics

This section summarizes the available

information on the absorption,

distribution, metabolism, and

elimination of the compound(s) in

humans and experimental mammals.

Conceptually, such information is useful

in validation of inter- and intraspecies

extrapolations, and in characterizing the

modes of toxic 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 of chemicals in the human

population. Distribution data which

suggest target organs or tissues are

desirable for interspecies comparison

techniques. In terms of the derivation of

criteria, pharmacokinetic data are

essential to estimate equivalent oral

doses based on data from inhalation or

other routes of exposure.

C. Effects

This section summarizes information

on biological effects in both humans and

experimental mammals resulting in:

acute, subacute, and chronic toxicity,

synergism and/or antagonism,

teratogenicity, mutagenicity, or

carcinogenicity.

The major goal of this section is to

survey the suitability of the data for use

in assessment of hazard 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 of such data in generating water

quality criteria. These problems, as

summarized by the National Academy

of Sciences (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. If exposure 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 of a new

material can sometimes be useful, but

when latent periods are variable and

times of introduction and removal of

materials overlap, historical data on

chronic effects are usually

unsatisfactory.

3. It is 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 if the

population is large.

5. Interactions in a "nature-designed"

experiment usually cannot be

controlled.

Although these problems often

prevent the use of epidemiological 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 of animal-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 a non-threshold model is made

after evaluating all available

information on carcinogenicity and

supportive information on mutagenicity.

The approach and conditions for the

qualitative decision of carcinogenicity

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). It is 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 in humans or experimental

animals is not identified as a

carcinogen, even though Its metabolites

or close structural analogues might

induce a carcinogenic response or it was

shown to be mutagenic in an in vitro

system.

It is 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 of carcinogenicity 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

of a criterion for human consumption

from these studies in not valid,

regardless of the qualitative outcome. In

addition, there are 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 or humans via

ingesting their soluble salts.

Nevertheless, a non-threshold criterion

is developed for beryllium because

tumors have been produced 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 of tumors

at sites distant resulting from

administration of nickel solutions by

either ingestion or injection.

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

of human exposure, preference is given

to chronic studies involving oral

exposures in water or diet over a

significant portion of the 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 of environmental

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 of the compounds evaluated

thus far, teratogenicity and reproductive

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

I I II I I

79349
Federal 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 of the

life span of the test organism, e.g., 90

days for the rat with an average life

span of 30 months. Such stildies are

useful in establishing the nature of the

compound's toxic effects and other

parameters of compound toxicity, such

as target organ effects, metabolic

behavior, physiological/biochemical

effects,'and patterns of retention and

tissue distribution. The utility of acute

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 of administration, 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

of absorption.

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 of Governmental 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

of Sciences in Drinking Water and

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 of current levels of human

exposure, presented in this section,

-specifically address the suitability of the

data to derive water quality criteria. The

identification of special groups at risk,

either because of geographical 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 of water-quality criteria.

The basis for the recommended

criteria section summarizes and

qualifies all of.the data used in

developing the criteria.

IV. Guidelines for Criteria Derivation

The derivation of water quality

criteria from laboratory animal toxicity

duta is essentially a two-step procedure:

First, a total daily intake for humans

must be estimated which establishes

either a defined level of risk for nonthreshold effects or a no-effect leiel for

threshold effects. Secondly, assumptions

must be made about the contribution of

contaminated water and the

consumption of fish/shellfish to the total

daily intake of the chemical. These

estimates are then used to establish the

tolerable daily intake and consequently

the water quality criterion.

A. Non-Threshold Effects

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

of a criterion, the water concentration

which is estimated to cause a lifetime

carcinogenic risk of 10- is -determined.

Thb lifetime carcinogenicity risk is the

probability that a person wohld get

cancer sometime in his or her life

assuming continuous exposure to the

compound. The water concentration is

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 of two types: (1) lifetime

animal studies, and (2) human studies

where excess cancer risk has-been

associated with exposure to the agent.

In animal studies it is assumed, unless.

evidence exists to the contrary, that if 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 of Model.

There is no really solid scientific basis

for any mathematical extrapolation

model which relates carcinogen

exposure to cancer risks at the

-extremely low levels of concentration

that must be dealt with in evaluating the

environmental hazards. For practical

reasons, such low levels of risk cannot

be measured directly either using animal

experiments or epidemiologic studies.

We must, therefore, depend on our

current understanding of the

mechanisms of carcinogenesis for

guidance as to which risk model to use.

At the present time, the dominant view

of the carcinogenic process involves the

concept that 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 of biological

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 of chemicals that this type of'

dose-response model is the appropriate

one to use. This is particularly true at

the lower end of the dose-response

curve; at higher doses, there can be an

upward curvature, probably reflecting

the effects of multistage processes on

the mutagenic response. The linear nonthreshold dose-response relationship Is

also consistent with the relatively few

epidemiological studies of cancer

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 that is consistent with the

linear non-threshold hypothesis (e.g.,

liver tumors induced in mice by 2-

.acetylaminofluorene in the large scale

EDo1 study at the National Center of

Toxicological Research, and the

initiation stage of the two-stage

carcinogenesis model in the rat liver and

the mouse skin).

Because it has the best, albeit limited,

scientific basis of any of the current

mathematical extrapolation models, the

linear non-threshold model has been

adopted as the primary basis for risk

extrapolation to low levels of the 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 risk is not

likely to be higher than the estimate, but

it could be smaller. I

79350
Federal Register / VoL 45, No. 231 / Friday, November 28, 1980 / Notices

The mathematical formulation chosen

to describe the linear, non-threshold

dose-response relationship at low 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 largest possible 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 this reason, it may be

called a 'linearized" multistage model.

2. Procedure of Low-Dose

Extrapolation Based on Animal

Carcinogenicity Data.

A. Description of the Extrapolation

Model

Let P(d) represent the lifetime risk

(probability) of cancer at dose d. The

multistage model has 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 over background rate at

dose d.

The point estimate of the coefficients

04, i=O, 1, 2 .... ,k, and consequently

the extra risk function A(d) at any given

dose d, is calculated by maximizing the

likelihood function of the data.

The point estimate and the 95 percent

upper confidence limit of the extra risk

A(d) are calculated by using the

computer program GLOBAL 79

developed by Crump and Watson (1979).

Upper 95 percent confidence limits on

the extra risk and lower 95 percent

confidence limits on the dose producing

a given risk are determined from a 95

percent upper confidence limit, q1*, on

parameter qi. Whenever q,=*0, at low

doses extra risk A(d) has approximately

the form A(d)=q , xd. Therefore, qxd

is a 95 percent upper confidence limit on

the extra risk and R/qi* is a 95 percent

lower confidence limit on the dose

producing an extra risk of R. Let L. be

the maximum value of the log-likelihood

function. The upper 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

resulting maximum value of the loglikelihood 14 satisfies the equation

2(L.-I,)=2.7G554

where 2.70654 is the cumulative 90

percent point of the chi-square

distribution with one degree of freedom,

which corresponds to a 95 percent upper

limit (one-sided). This approach of

computing the upper confidence 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 at low 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 of the

chemical in inducing cancer at low

doses.

In fitting the dose-response model, the

number of terms in the polynomial g Is

chosen equal to (h-i]. where h is the

number of dose groups in the

experiment. including the control group.

Whenever the multistage model does

not fit the data sufficiently, data at the

highest dose is deleted and the model is

refitted to the rest of the data. This is

continued until an acceptable fit to the

data is obtained. To determine whether

or not a fit is acceptable, the chi-square

statistic:

h

X2= (Xi - NiPi)2

NiPi (1 - Pi)

is calculated, where N, is the number of

animals in the i'b dose group, X1 is the

number of animals in the i" dose group

with a tumor response, Pi is the

probability of a response in the ill dose

group estimated by fitting the multistage

model to the data, and h Is the number

of remaining groups.

The fit is determined to be

unacceptable whenever chi-square (XJ

is larger than the cumulative 99 percent

point of the chi-square distribution with

f degrees of freedom, where f equalsthe

number of dose groups minus the

number of non-zero multistage

coefficients.

3. Selection and Form of Data 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 of exposure are

available. A choice must be made as to

which of the data sets from several

studies are to be used in the model. It is

also necessary to correct for metabolism

differences between species and for

differences in absorption via different

routes of administration. The

procedures, listed below, used in

evaluating these data are consistent

with the estimate of a maximum-likelyrisk.

a. The tumor incidence data are

separated according to organ sites or

tumor types. The set data (i.e., dose and

tumor incidence) used in the model is

set where the incidence is statistically

significantly higher than the control for

at least one test dose 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 of lifetime

carcinogenic risk q1 * is selected in most

cases. However, efforts are made to

exclude data sets which produce

spuriously high risk estimates because

of a small number of animals. That is, if

two sets of data show a similar doseresponse relationship and one has a

very small sample size. the set of data

which has the larger sample size is

selected for calculating the carcinogenic

potency.

b. If there are two or more data sets of

comparable size which are identical

with respect to species, strain. sex. and

tumor sites, the geometric mean of qt*,

estimated from each of these data sets is

used for risk assessment. The geometric

mean of numbers A, Az ... , A. is

defined as (AzxAX ... XA.)J"

c. If sufficient data exist for two or

more significant tumor sites in the same

study, the number of animals with at

least one of the specific tumor sites

under consideration is used as incidence

data in the model.

d. Following the suggestion of Mantel

and Schneiderman (1975). we assume

that mg/surface area/day is an

equivalent dose between species. Since

to a close approximation the surface

area is proportional to the %rds power

of the weight as would be the case for a

perfect sphere, the exposure in mg/%rds

power of the 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 of exposure

m=average dose per day in mg during administration of the agent (Le, during 1])

W= averge weight of the experimental animal

Then, the lifetime average exposure is

d l e x m

Le x W2 /3

Often exposures are not given in units

of mg/day, and it becomes necessary to

convert the given exposures into mng/

day. For example, in most feeding

studies, exposure Is expressed as ppm in

the diet. In this case the exposure (mg/

day) is derived by: m=ppm x F x r

79351
Federal Register / Vol. 45, No. 231 / Friday, November 28, 1980 / Notices

where ppm is parts per million of the

carcinogenic agent in the diet, F is the

weight of the food consumed per day in

kgms, and r is the absortion fraction.

In the absence of any data to the

contrary, r is-assumed to be one. For a

uniform diet the weight of the food

consumed is proportional to the calories

required, which, in turn, is proportional

to the surface' area or the %rds power of

the weight, so that: mappmXW2/1 Xr or

in ppm

As a result, ppm in the diet is often

assumed to be an equivalent exposure

between species. However, we feel that

this is not justified since the calories/kg

of food is significantly different in the

diet of man vs. laboratory animals,

primarily due to moisture content

differences. Instead, we use an

empirically derived food factor, f=F/W,

which is the fraction of a 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/ 3 is

m = ppm ix F

r x W2 / 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

rW2 / 3

When exposure is via inhalation, the

calculation of dose 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

of air 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 of absorption of these

agents is expected to be proportional to

metabolic rate, which in turn is

proportional to the rate of oxygen

consumption, which in turn is a function

of surface area.

Case 1

Agents that are in the form of

particulate iatter or virtually

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 in m, v is

mg/m a of the 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 of other weights, W, (expressed in

kg], the surface area proportionality can

be used to-determine breathing rates (in

ms/day) as follows:

For mice, I=0.0345 (W/0.025) 2/ 1!10/

day -

For rats, I=0.105 (W/0.113) 21Sms/day For humans, the values of 20 m3/day a

is adopted as a standard breathing rate

(ICRP, 1977).

The equivalent exposure in mg/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 per kg per day,

i=I/W based 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, for particulates or completely

absorbed gases, the equivalent exposure

inmg/W2'3 is:

mr Ivr = iWvr = jw1/3 vr

T21_3 721/3 7213

In the absence of empirical 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 of partially

soluble vapois is proportional to the Qt'

consumption which in turn is

proportional to W 213 and to the

solubility of gas in body fluids, which

can be expressed as an absorption

coefficient r for the gas. Therefore, when

expressing the 06 consumption as O2=k

W2/3 , where k is a constant independent

a From "Recommendation of the International

Commission on Radiological Protection." page 9, the

iverage breathing rate is 107

cm~per'8-hour work

day and 2X0 7 cm3 in 24 hours.

of species, it follows that m=k W21 x V

x r or

d m kvr

w2 13'

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 3 in 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" of anesthesia, 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 or vice-versa, the assumption

is made, unless there is pharmacokineflo

evidence to the contrary, that absorption

is equal by either exposure route,

e. If the duration of experiment (L.) Is

less than the natural life span of the4est

animal (L), the slope qa*, or more

generally the exponent g(d), is increased

by multiplying a factor (L/L.)3 . We

assume that if the average dose, d, is

continued, the age specific rate of

cancer will continue to increase as a

constant function of the background

rate. The age specific rates for humans

,increase at least by the 2nd power of the

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 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 of age. As a result, if

the slope q1 * [or g(d)] is calculated at

age L0, we would expect that if the

experiment had been 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

model proposed by Cox (1972) and the

time-to-tumor model considered by

Crump, et al. (1977) vhere the

probability of cancer at age t and dose d

is given by P(d,t)=1-exp[-f(t)Xg(d)] 4. Calculation of Carcinogenic Potency

Based on Human Data. If human

epidemiology studies and sufficiently

valid exposure information are available

for the compound, they are always used

in some way. If they show a

carcinogenic effect, the data are

analyzed to give an estimate of the

linear dependence of cancer rates on

lifetime average dose, which Is

equivalent to the factor q*. If they show

79352
Federal Register / VoL 45, No. 231 / Fridak. November 28, 1980 / Notices

no carcinogenic effect when positive

animal evidence is available, then it is

assumed that a risk does exist but it is

smaller than could have been observed

in the epidemiologic study, and an upper

limit of the cancer incidence is

calculated assuming hypothetically that

the true incidence is just below the level

of detection 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 of individuals

compared to the control group. In the

analysis of this data, it is assumed that

the excess risk, or relative 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 an unusually

well-documented human study, the

confidence limit for the excess risk is

not calculated, due to the difficulty in

accounting for the uncertainty inherent

in the data (exposure and cancer

response).

5. Calculation of Water Quality

Criteria. After the value of qi* in [mg/

kg/day)-I has been determined, the

lifetime risk, P, from an average daily

exposure of x mg/kg/day is found from

the equation P=q,*x. Therefore, if the

lifetime risk is set at P=10-5 for

calculation purposes, the intake, I, in

mg/day for a 70 kg person can be found

by the equation: I=70X10-S/q*

The intake of the agent from ambient

water is assumed to come from two

sources: (1) drinking an average of 2

liters of water per day, and (2) ingesting

an average of 6.5 grams of fish per day.

Because of accumulation of residues in

fish, the amount of the pollutant in fish

(mg/kg of edible fish) is equal to a factor

R times the water concentration (mg/kg

of water). Therefore, the total intake I

can be written as sum of two 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 kg person is calculated by the

formula:

$C = 70 x 10-5 C = o1*(2 + 0.0065 R)

B. Threshold Effects

1. Use of Animal Toxicity Data (Oral).

n developing guidelines for deriving

criteria based on noncarcinogenic

responses, five types of response levels

are considered:

NOEL-No-Observed-Effect.Level

NOAEL--No-Observed-Adverse.Effect.Level

LOEL--Lowest-Observed.Effect.Level

LOAEL--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, or which reduce an

organism's ability to respond to an

additional challenge.

One of the major problems

encountered in consideration of these

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 due to

increasingly sophisticated testing

protocols, more subtle responses are

being identified, resulting in a need for

judgment regarding the exact definition

of adversity.

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, If a single experiment yields a

NOEL a NOAEL'a LOAEL and a

clearly defined FEL in relatively closely

spaced doses, the threshold region has

been relatively well defined. such data

are very useful for the purpose of

deriving a criterion. On the other hand. a

clearly defined FEL has little utility in

establishing criteria when it stands

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 of its proximity to the LOEL.

since a free-standing NOEL may be

many orders of magnitude 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 FEL is unsuitable

for the derivation of criteria.

b. A free-standing NOEL is unsuitable

for the derivation of criteria. If multiple

NOELs are available without additional

data on LOELs, NOAELs, or LOAELs,

the highest NOEL should be used to

derive a criterion.

c. A NOAEL, LOEL, or LOAEL can be

suitable for criteria derivation. A welldefined NOAEL from a chronic (at least

90-day) study may be used directly,

applying the appropriate uncertainty

factor. For a LOEL, a judgment 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 of the

additional uncertainty factor is

judgmental and should lie in the range of

1 to 10. Caution must be exercised not to

substitute "Frank-Effect-Levels" for

"Lowest-Observable-Adverse-EffectLevels".

d. If for 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

and justification of uncertainty factors

are critical. The basic definition and

guidelines for using uncertainty factors

has been given by the National

Academy of Sciences (1977). "Safety

Factor" or "Uncertainty Factor" is

defived as a number that reflects the

degree or amount of uncertainty that

must be considered when experimental

data in animals are extrapolated to man.

When the quality and quantity of

experimental data are satisfactory, a

low uncertainty factor is used; when

data is judged to be inadequate or

equivocal, a larger uncertainty factor is

used. The following general guidelines

have been adopted in establishing the

uncertainty factors:

a. Valid experimental results from

studies on prolonged ingestion by man.

with no indication of carcinogenicity.

Uncertainty Factor=10

b. Experimental results of studies of

human ingestion not available or scanty

(e.g., acute exposure only) with valid

results of long-term feeding studies on

experimental animals, or in the absence

of human studies, valid animal studies

on one or more species. No indication of

carcinogenicity. Uncertainty Factor=100

c. No long-term or acute human data.

Scanty results on experimental animals

with no indication of carcinogenicity.

Uncertainty Factor=1,000

Considerable judgment must be 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

factor is used. Such an intermediate

uncertainty factor may be developed

based on a logarithmic scale (e.g., 33,

being halfway between 10 and 100 on a

logarithmic scale).

In determining the appropriate use of

the uncertainty factors, the phrase "no

79353
Federal Register / Vol. 45, No. 231 / Friday, November 28, 1980 / Notices

indication of carcinogenicity" 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

numerigal criteria nor to rule out the

uncertainty factor approach.

Because of the high degree of

judgment involved in the selection of a

safety factor, the criterion derivation

section of each document should

provide a detailed discussion and

justification for both the selection of the

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 of animals, 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 of Acceptable Daily Intake

(ADI). For carcinogens, the assumption

of low dose linearity precludes the

necessity for defining total exposure in

the estimation of increased 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 1 is assumed daily water

consumption, 0.0065 kg is assumed daily

fish consumption, R is bioconcentration

factor in units of 1/kg, DT is estimated

non-fish dietary intake, and IN is

estimated daily intake by inhalation.

If estimates of IN 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 of data 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 of water quality

criteria.

3. Use of Threshold Limit Values 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 set by 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 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

each TLV 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 of the routes of

exposure (i.e., oral or inhalation).

If the TLV or similar value is

recommended as the basis of the

criterion, consideration of the above

points is explicitly stated in the criterion

derivation section of the document.

Particular emphasis is placed on the

quality of the TLV relative to the

available toxicity data that normally is

given priority over TLVs or similar

established values. If the TLV can be

justified as the basis for the cirterion,

then the problems associated with the

estimation.of acceptablezoral doses from

inhalation data mustbe addressed.

Estimating equivalencies of doseresponse relationships from one route of

exposure to another introduces an

additional element of uncertainty in the

derivation of criteria. Consequently,

whenever possible, ambient water

quality criteria should be based on data

involving oral exposures. if oral 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

metabolic pathways are equivalent to

those following oral ingestion or

inhalation. Given that sufficient

pharmacokinetic data are available, the

use of accepted pharmacokinetic models

provides the most satisfactory approach

for dose conversions. However, if

available phiarmacokinetic data are

marginal or of questionable quality,

pharmacokinetic modeling is

inappropriate. The Stokinger and Woodward (1958)

approach, or similar models based on

assumptions of breathing rate and

absorption efficiency, represents

possible alternatives when data are not

sufficient to justify pharmacokineto

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 of the

assumptions and the basis of each

assumption must be clearly stated in the

derivation of the criterion.

The use of data 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. If simplifying

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.

If the 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 air in mg/m 3

DE=Duration of exposure In hours per day

d=5 days/7 days

AA=Efficiency of absorption from air

Ao=Efficiency of absorption from oral

exposure

SF= Safety factor following guidelines given

above

BR=Amount of air breathed per day; assume 10 m3

For deriving an ADI from animal

toxicity data, the equation Is:

ADI=CAXDExdxAAXBRX70 kg/

(BWA X Ao X SF) where:

ADI=Acceptable daily intake in mg

CA= Concentration in air In mg/m 3

DE=Duration of exposure in hours per day

d=Number of days exposed/number of days

observed

AA=Efficiency of absorption from air

BR=Volume of air breathed per day in m3

70 kg=Assumed human body weight

BWA=Body weight of experimental animals

in kg

A354

79a54
Federal Register / Vol. 46. No. 231 / Friday, November 28, 1980 / Notices

AO=Efficieay of absorption from oral

exposure

SF=Safety factor following 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 of toxicity

based criterion.

C. Organoleptic Criteria

Organoleptic criteria define

concentrations of materials which

impart undesirable taste and/or odor to

water. In developing and utilizing such

criteria two factors must be appreciated:

the limitations of most 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

of test 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 of toxicologic

data used in establishing other criteria.

Consequently, a critical evaluation of

the available organoleptic data must be

made and the selection of the 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 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 of organoleptic

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 of information.

In criteria summaries involving both

types of data, the following format is

usech

For comparison purposes, two approaches were used to derive criterion levels for .Based on available toxicity data.

for the protection of public health the derived

level is -. Using available organoleptic

data. for controlling undesirable taste and

odor quality of ambient 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 of this

compound.

D. Criteria for Chemical Classes

A chemical class is broadly defined as

any group of chemical 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 of risk/safety which applies to

more than one member of a 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 of environmental

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 of similarity among

chemicals within a class to justify a

class criterion applicable to some of all

members of a class.

The development of a class criterion

takes into consideration the following:

1. A detailed review of the chemical and

physical properties of chemicals within the

group should be made. A close relationship

within the class with respect to chemical

acitivity would sugest a similar potential to

reach common biological sites within tissues.

Likewise, similar lipid solubili ties would

suggest the possibility of comparable

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 one 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 to other 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 of a criterion to other members

of that class is not appropriate.

4. Additional support for the validity of

extrapolation of a criterion to other members

of a class could be provided by evidence of

similar metabolic and pharmacokinetic data

fur some members of the class.

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 of data (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

of subclassification, each

subclassification must be explicitly

justified by the available data.

Class criteria, if properly derived and

supported, can constitute valid scientific

assessments of potential risk/safety.

Conversely, the development of a class

criterion from an insufficient data base

can lead to serious errors in

underestimating or overestimating risk/

safety and should be rigorously avoided.

Although scientific judgment has a

proper role in the development of class

criteria, such criteria are uieful and

defensible only if they are based on

adequate data and scientific reasoning.

The definition of sufficient 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 of the perceived 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

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

limitations of the proposed criterion as

well as the type of data 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 of primary concern; e.g., the.

availability of a good subchronic study

would not necessarily result in the

abandonment of a class criterion based

on potential carcinogenicity. The inability to derive a valid class

criterion does not, and should not,

preclude regulation of a compound or

group of compounds 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. 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 of potential 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 of the 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 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 of primary concern.

The difference between the RDA or

MDR and the daily doses causing a

specified risk level for carcinogens or

ADIs 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 decisiorto use either the MDR or

RDA is guided by the spread of the

doses and the quality of the essentiality and toxicity estimates.

The modification of criteria by

consideration of essentiality must take

into accountall routes of exposure. If

water is a significant source of the'MDR

or RDA, the criterion must allow for

attainment of essential intake.

Conversely, even when essentiality may

be attained from nonwater sources,

standard criteria derivation methods

may be adjusted if the derived criterion

represents a small fraction of the ADI or

MDR. On a case-by-case basis, the

modification in the use of the guidelines

may include the use of different 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

standards represent not only a critical

assessment of literature, but also a body

of human 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 which has- . accumulated since the standards have

been established. There are several

instances where the peer review process

recommended usage of the present

drinking water standards.

* Bibliography

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. 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. (In press).

Crump, K.S. and W.W. Watson. 1979.

GLOBAL 79. A FORTRAN 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 of hypotheses 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., et al. 1977. Introduction to

Anesthesia, the Principles of Safe 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 of Risks. UICC Monograph Series,

Vol. 7. Springer Verlag, New York.

Federation of American Societies for

Experimental Biology. 1974, Biological Data

* Books, 2nd ed. Vol. I. P.L. Altman and D.S.

Dittment (eds.) Federation of American

Societies for Experimental Biology. Bethesda,

Maryland. Library of Congress No. 72-87738.

Guess, H., et al. 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 of Commercial and Industrial

.Chemicals. In: M.A. Mehimar, et al. (eds.),

Advances in Modem Toxicology, Vol. 1, Part

1. Wiley and Sons, New York, p. 61

National Academy of Sciences. 1977.

Drinking Water and Health. National

Academy of Sciences, 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. National Interim Primary

Drinking Water Regulations. Office of Water

Supply. EPA 570/9-76-003. 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,

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 I School, 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 of Rochester,

Rochester, NY

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Federal 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 of Michigan. 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 of CincinnatL

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 of Cincinnati.

Cincinnati, OH

Fred Passman, Energy Resources Co.. Inc.

Cambridge, MA

John F. 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

Roll Hartung. University of Michigan. Ann

Arbor, MI

Rudolf Jaeger, 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

Leland McCabe, HERL-Cin, U.S.

Environmental Protection Agency

Myron Mehiman, Mobil Oil Corporation, New

York, NY

Jean Munson. University of Cincinnati,

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 for Deriving Water Quality

Criteria for the Protection of Aquatic

Life and Its Uses

Introduction

Two versions of the 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 of the 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 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 of the technical

content of the 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 simply by

not publishing any version of the

Guidelines for comment until March 15.

1979, but this would have greatly

reduced the length of 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 "Red Book" were

upheld in court and probably should still

be used.

Response-The procedures used in

the Guidelines are similar to some of the

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 of a minimum data base.

2. Comment-Criteria should be

compilations of critically reviewed data

with no synthesis or interpretation.

Response-Neither P.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, inter al'a,

recommended maximum permissible

concentrations". Adequate precedents

have been set in the "Green Book',

"Blue Book', and "Red Book" for the

form of criteria used in the Guidelines.

3. Comment-The Guidelines and"

criteria should be developed by a

consensus of aquatic 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 of the 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

reached by 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 of presenting

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

will be revised to take into account new

data, concepts, and ideas.

5. Comment-The objectives, purpose,

and limitations of the Guidelines should

be stated.

Response-The introductory portion

of the Guidelines has been 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 and provide additional details. It

is not possible to provide explicit details

on all items; in some areas only general

guidance can be provided at this time.

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