United Stater Office of Water 

Environmental Protection Regulatmr and Standards 

Agency Warhmgton. DC 20460 May 1, 1986 

e,Ep', Water EPA 440/5-86-001 

QUALlTY CRITERIA 

for 

WAm 

1986 
- TO INTERESTED PARTIES 

Section 304(a)(l) of the Clean Water Act (33 U.S.C. 

1314(a) (1) requires the Environmental Protection Agency (EPA) to 

pub1 ish and periodically update ambient water qua1 ity criteria. 

These criteria are to accurately reflect the latest scientific 

knowledge (a) on the kind and extent of all identifiable effects 

on health and welfare including, but not limited to, plankton, 

fish shellfish, wildlife, plant life, shorelines, beaches, 

aesthetics, and recreation which may be expected from the 

presence of pollutants in any body of water including ground 

water; (b) on the concentration and dispersal of pollutants, or 

their byproducts, through biological, physical, and chemical 

processes; and (c) on the effects of pollutants on biological 

community diversity, productivity, and stability, including 

information on the factors affecting rates of eutrophication and 

organic and inorganic sedimentation for varying types of 

receiving waters. These criteria are not rules and they do not 

have regulatory impact. Rather, these criteria present 

scientific data and guidance of the environmental effects of 

pollutants which can be useful to derive regulatory requirements 

based on considerations of water quality impacts. When 

additional data has become available, these summaries have been 

updated to reflect the latest Agency recommendations on 

acceptable limits for aquatic life and human health protection. 

Periodically EPA and its predecessor agencies has issued 

ambient water quality criteria, beginning in 1968 with the "Green 

Book" followed by the 1973 publication of the "Blue Book" (Water 

Quality Criteria 1972). In 1976, the "Red Book" (Quality 

For aalc by the Svpmlendsnt of Documents, US Mwmmsnt RlnUng ORW 

Waahpton. DC M02 
Criteria for Water) was published. On November 28, 1980 (45 FR 

79318), and February 15, 1984 (49 FR 5831), EPA announced through 

Federal - Register notices, the publication of 65 individual 

ambient water quality criteria documents €or pollutants listed as 

toxic under section 307(a)(l) of the Clean Water Act. on July 

29, 1985 (50 FR 30784), EPA published additional water quality 

criteria documents. 

The development and publication of ambient water quality 

criteria has been pursued over the past 10 years and is an 

ongoing process. EPA expects to publish about 10 final criteria 

documents each year. Some of these will update and revise 

existing criteria recommendations and others will be issued for 

the first time. 

In a continuing effort to provide those who use EPA’S water 

quality and human health criteria with up-to-date criteria values 

and associated information, this document -Qualm - Criteria __ €or 

Water - 1986 was assembled. This document includes summaries of 

all the contaminants for which EPA has developed criteria recommendations (Appendix A-C) . The appropriate appendix is 

identified at the end of each summary. A more detailed 

description of these procedures can be found in the appropriate 

Appendix. Copies of this document can be obtained by contacting 

the U.S. Government Printing Office at: 

U.S. Government Printing Office 

Superintendent of Documents 

N. Capitol and H Street N.W. 

Washington, D.C. 20401 

A fee is charged for this document. 

Copies of the complete background ambient water quality 
criteria documents containing all the data used to develop the 

criteria recommendations summarized herein and the “Red Bookt8, 

including complete bibliographies are available only from: 

0 

National Technical Information Service 

5285 Port Royal Road 

Springfield, VA 22161 

Telephone: (703) 487-4650 

The NTIS order numbers for the criteria documents can be found in 

the Index. A fee is charged for copies of these documents. 

As new criteria are developed and existing criteria revised, 

updated criteria summaries will be made available once a year to 

those who purchase this document through the U.S, Government 

Printing office. You will automatically be placed on the mailing 

list to receive annual updates. The cost for receiving annual 

updates is included in the purchase price of the document. 

-Quality - Criteria - for Water, 1986 is designed to be easily 

updated to reflect EPA‘s continuing work to present the latest 

scientific information and practices, Our planned schedule €or 

future criteria development in the next few years is attached for 

your information. 

a 

The Agency is current1 y developing Acceptable Daily Intake 

(ADI) or Verified Reference Dose (RfD) values on a number of 

chemicals for Agency-wide use. Based upon this new analysis the 

values have changed significantly for 5 chemicals from those used 

in the original human health criteria calculation done in 1980. 

The chemicals affected are as follows: 

0 
chemical . 1980 WQC 

1. cyanide 200 ug/L 

2. Ethylbenzene 1.4 mg/L 

4. Phenol 3.5 mg/L 

3. Nitrobenzene 19.8 mg/L 

5. Toluene 14.3 mg/L 

Draft RfD 

.02 mg/kg/day 

.Q1 mg/kg/day 

.Q005 mg/kg/day 

0.1 mg/kg/day 

0.3 mg/kg/day 

FOR FORTHER INFORMATION CONTACT: 

Dr. Frank Gostomski at the above address or by phoning (202) 245- 

3030. 

It is EPA's goal to continue to develop and make available 

ambient water quality criteria reflecting the latest scientific 

practices and information. In this way we can continue to 

improve and protect the quality of the Nation's waters. 

James M. Conlon 

i/ and Standards 
DRAFT CRITERIA DOCUMENTS TO BE PROPOSED 

-- LATE FY 86/EARLY -- FY 87 

0 

Diethyhexylphthalate 

1,2,4, Trichlorobenzene 

Silver 

Phenanthrene 

2,4,5, Trichlorophenol 

Organotins 

Tributyltin 

Selenium (no saltwater criteria) 

Hexachlorobenzene 

Antimony 111 

Acrolein (no saltwater criteria) 

-LATE FY - 87/EARLY 

Thallium (no saltwater criteria) 

Tetrachloroethylene (no saltwater criteria) 

Phenol 

Toluene 

Chloroform (no saltwater criteria) 

'imaline 

w Acrvlontrile Hexachlorocyclopentadiene (no saltwater criteria) 

Dimethylphenol 

Hexachlorobutadiene (no saltwater criteria) 

- Both lists will incorporate aquatic and human health values. 

- All above are toxic pollutants except for organotins and 

analine which are non-conventionals. 

INDEX 

INTRODUCTION 

SUMMARY CHART 

.,' 

Acenaphthene 

Acrolein 

Acrylonitrile 

Aesthetics 

Alkalinity 

Aldrin/Dieldrin 

Ammonia 

Antimony 

Arsenic 

Asbestos 

Bacteria 

Barium 

Benzene 

Benzidine 

Beryllium 

Boron 

Cadmium 

Carbon Tetrachloride 

Chlordane 

Chlorinated Benzenes 

Chlorinated Ethanes 

Chlorinated Naphthalenes 

Chlorine 

Chlorinated Phenols 

Chloroalkyl Ethers 

Chlorof om 

Chlorophenoxy Herbicides 

Chromium 

2-Chlorophenol 

Color 

Copper 

Cyanide 

DDT and Metabolites 

Demeton 

Dichlorobenzenes 

Dichlorobenzidine 

Dichloroethylenes 

2,4, - Dichlorophenol 

Dichloropropanes/Dichloropropenes 

2,4, - Dimethylphenol 

Dinitrotoluene 

Diphenylhydrazine 

Endosulfan 

Endrin 

Ethylbenzene 

Fluoranthene 

Gasses, Total Dissolved 

Guthion 

PB-263943 

PB 81-117293 

PB 81-117343 

PB 81-117350 

PB-2 6394 3 

PB 85-227031 

PB 81-117376 

PB 81-117384 

PB 81-117392 

PB 81-117400 

PB 81-117426 

PB 85-227429 

PB 81-117434 

PB 81-117418 

PB 81-117442 

PB-2 63 94 3 

PB 85-227478 

PB 81-117459 

PB-2 63 94 3 

PB 85-227023 

PB 85-227460 

PB 81-117491 

PB-263943 

PB 81-117509 

PB 81-117517 

PB 81-117525 

PB 81-117533 

PB 81-117541 

PB 81-117558 

PB 81-117566 

PB 81-117731 

PB 81-117582 

PB 81-117590 

PB 81-117608 

PB-2 63 943 

PJ3-2 639 4 3 

PB 81-117574 

-- NTIS No. 

PB 81-117269 

PB 81-117277 

PB 81-117285 

FB-2 63 94 3 

PB-2 63 94 3 

PB 81-117301 

PB 85-227114 

PB 81-117319 

PB 85-227445 

PB 81-117335 

PB 86-158-045 h PB-263943 
Haloethers 

Halomethanes 

Hardness 

Heptachlor 

Hexachlorobutadiene 

Hexachlorocyclohexane 

Hexachlorocyclopentadiene 

Iron 

Isophorone 

Lead 

Malathion 

Manganese 

Mercury 

Methoxychlor 

Mirex 

Naphthalene 

Nickel 

Nitrates, Nitrites 

Nitrobenzene 

Nitrophenols 

Nitrosamines 

Oil and Grease 

Oxygen, Dissolved 

Parathion 

Pentachlorophenol 

Ph 

Phenol 

Phosphorus 

Phthalate Esters 

Polychlorinated Biphenyls 

Polynuclear Aromatic Hydrocarbons 

Selenium 

Silver 

Solids (Dissolved) h Salinity 

Solids (Suspended) & Turbidity 

Sulfides, Hydrogen Sulfide 

Taintina Substances 

PB 81-117616 

PB 81-117624 

PB-263943 

PB 81-117632 

PE 81-117640 

PB 81-117657 

PB 81-117665 

PB-263943 

PB 81-117673 

PB 85-227437 

PB-263943 

PB-263943 

PB 85-227452 

PB-2 63 943 

PB-263943 

PB 81-117707 

PB 81-117715 

PB-263943 

PB 81-117723 

PB 81-117749 

PB 81-117756 

PB-263943 

PB 86-208253 

PB-263943 

PB 81-117764 

PB-263943 

PB 81-117772 

PB-263943 

PB 81-117780 

PB 81-117798 

PB 81-117806 

PB 81-117814 

PB 81-117822 

PB-263943 

PB-263943 

PB-263943 

PB-263943 

Temperagure PB-263943 

Tetrachloroethylene PB 81-117830 

Thallium PB 81-117848 

Toluene PB 81-117855 

Toxaphene PB 81-117863 

Trichloroethylene PB 81-117871 

Vinyl Chloride PB 81-117889 

Zinc PB 81-117897 

2,3,7,8-Tetrachlorodibenzo-p-dioxin EPA # 440/5-84-007 

APPENDIX A 

APPENDIX B 

APPENDIX C 

BIBLIOGRAPHY 

Methodology for Developing Criteria 

Methodology for Developing Criteria 

Methodology for Developing Criteria 
I SUMMARY CHART 1 
0 CRITERIA: 

ACENAPHTHENE 

Aquatic Life 

The available data for acenaphthene indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 1,700 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of acenaphthene to 

sensitive freshwater aquatic animals but toxicity to freshwater 

algae occur at concentrations as low as 520 ug/L. 

The available data for acenaphthene indicate that acute and 

chronic toxicity to saltwater aquatic life occurs at 

concentrations as low as 970 and 710 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. Toxicity to algae occurs at 

concentrations as low as 500 ug/L. 

Human Health 

Sufficient data are not available for acenaphthene to derive 

a level which would protect against the potential toxicity of 

this compound. Using available organoleptic data, to 

control undesirable taste and odor quality of ambient water 

the estimated level is 0.02 mg/L. It should be recognized that 

organoleptic data, have limitations as a basis for establishing 

water quality criteria, and have no demonstrated relationship to 

potential adverse human health effects. 0 , (45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
ACROLEIN 

Aquatic Life 

The available data for acrolein indicate that acute and 

chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 68 and 21 ug/L, respectively, and would 

occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for acrolein indicate that acute toxicity 

to saltwater aquatic life occurs at concentrations as low as 55 

ug/L and would occur at lower concentrations among species that 

are more sensitive than those tested. No data are available 

concerning the chronic toxicity of acrolein to sensitive 

saltwater aquatic life. 0 

Human Health 

For the protection of human health from the toxic properties 

of acrolein ingested through contamins'ed aquatic organisms, the 

ambient water criterion is determined o be 320 ug/L. 

For the protection of human health from the toxic properties 

of acrolein ingested through contaminated aquatic organisms 

alone, the ambient water criterion is determined to be 780 

ug/L* 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 

0 , 
CRITERIA: 

ACRYMNITRILE 

Aquatic Life 

The available, data for acrylonitrile indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 7,550 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No definitive 

data are available concerning the chronic toxicity of 

acrylonitrile to sensitive freshwater aquatic life but mortality 

occurs at concentrations as low as 2,600 ug/L with a fish species 

exposed for 30 days. 

Only one saltwater species has been tested with acrylonitrile 

and no statement can be made concerning acute or chronic 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects resulting from exposure to acrylonitrile 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at and lo-’. 

The corresponding recommended criteria are 0.58 ug/L, 0.058 

ug/L, and 0.006 ug/L, respectively. If these estimates are made 

0 , for consumption of aquatic organisms only, excluding consumption 
of water, the levels are 6.5 ug/L, 0.65 ug/L, and 0.065 ug/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
AESTHETIC QUALITIES 

CRITERIA: 

a 

All waters free from substances attributable 

to wastewater or other discharges that: 

settle to form objectionable deposits; 

float as debris, scum, oil, or other 

matter to form nuisances; 

produce objectionable color, odor, taste, 

or turbidity; 

injure or are toxic or produce adverse 

physiological responses in humans, 

animals or plants: and, 

produce undesirable or nuisance aquatic 

life. 

RATIONALE: 

Aesthetic qualities of water address the general principles 

laid down in common law. They embody the beauty and quality of 

water and their concepts may vary within the minds of individuals 

encountering the waterway. A rationale for these qualities 

cannot be developed with quantifying definitions; however, 

decisions concerning such quality factors can portray the best in 

the public interest. 

_., 
Aesthetic qualities provide the general rules to protect 

water against environmental insults: they provide minimal freedom 

requirements from pollution; they are essential properties to 

protect the Nation's waterways. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
ALKALINITY 

CRITERION: 

20 mg/L or more as CaC03 freshwater aquatic life except where 

natural concentrations are less. 

INTRODUCTION: 

Alkalinity is the sum total of components in the water that 

tend to elevate the pH of the water above a value of about 4.5. 

It is measured by titration with standardized acid to a pH value 

of about 4.5 and it is expressed commonly as milligrams per liter 

of calcium carbonate. Alkalinity, therefore, is a measure of the 

buffering capacity of the water, and since pH has a direct effect 

on organisms as well as an indirect effect on the toxicity of 

certain other pollutants in the water, the buffering capacity is 

important to water quality. Examples of commonly occurring 

materials in natural waters that increase the alkalinity are 

carbonates, bicarbonates, phosphates and hydroxides. 

RATIONALE : 

The alkalinity of water used for municipal water supplies is 

important because it affects the amounts of chemicals that need 

to be added to accomplish calculation, softening and control of 

corrosion in distribution systems. The alkalinity of water 

assists in the neutralization of excess acid produced during the 

addition of such materials as aluminum sulfate during chemical 

coagulation. Waters having sufficient alkalinity do not have to 

be supplemented with artificially added materials to increase the 

i alkalinity. Alkalinity resulting from naturally occurring 
a materials such as carbonate and bicarbonate is not considered a 

health hazard in drinking water supplies, per se, and naturally 

occurring maximum levels up to approximately 400 mg/L as calcium 

carbonate are not considered a problem to human health (NAS, 

1974). 

Alkalinity is important for fish and other aquatic life in 

freshwater systems because it buffers pH changes that occur 

- naturally as a result of photosynthetic activity of the 

chlorophyll-bearing vegetation. Components of alkalinity such as 

carbonate and biocarbonate will complex some toxic heavy metals 

and reduce their toxicity markedly. For these reasons, the 

National Technical Advisory Committee (NATC, 1968) recommended a 

minimum alkalinity of 20 mg/L 'and the subsequent NAS Report 

(1974) recommended that natural alkalinity not be reduced by more 

than 25 percent but did not place an absolute minimal value for 

it. The use of the 25 present reduction avoids the problem of 

establishing standards on waters where natural alkalinity is at 

or below 20 mg/L. For such waters, alkalinity should not be 

further reduced. 

The NAS Report recommends that adequate amounts of alkalinity 

be maintained to buffer the pH within tolerable limits for marine 

waters. It has been noted as a correlation that productive 

waterfowl habitats are above 25 mg/L with higher alkalinities 

resulting in better waterfowl habitats (NATC, 1968). 
Excessive alkalinity can cause problems for swimmers by 

altering the pH of the lacrimal fluid around the eye, causing 

irritation. 

0 

For industrial water supplies, high alkalinity can be 

damaging to industries involved in food production, especially 

those in which acidity accounts for flavor and stability, such as 

the carbonated beverages. In other instances, alkalinity is 

desirable because water with a high alkalinity is much less 

corrosive. 

A brief summary of maximum alkalinities accepted as a source 

of raw water by industry is included in Table 1. The 

concentrations listed in the table are for water prior to 

treatment and thus are only desirable ranges and not critical 0 

ranges for industrial use. 

The effect of alkalinity in water used for irrigation may be 

important in some instances because it may indirectly increase 

the relative proportion of sodium in soil water. As an example, 

when bicarbonate concentrations are high, calcium and magnesium 

ions that are in solution precipitate as carbonates in the soil 

water gs the water becomes more concentrated through evaporation 

and transpiration. As the calcium and magnesium ions decrease in 

concentration, the percentage of sodium increases and results in 

soil and plant damage. Alkalinity may also lead to chlorosis in 

plants because it causes the iron to precipitate as a hydroxide 

(NAS, 1974). Hydroxyl ions react with available iron in the soil 
TABLE I* 

Maximum Alkalinity In Waters Used As A Source 

Of Supply Prior To Treatment 

Industry 

Alkalinity 

mg/L as CaC03 

- 

Steam generation boiler makeup....... ..... 350 

Steam generation cooling .................. 500 

Textile mill products ..................... 50-200 

Paper and allied products ................. 75-150 

Chemical and Allied Products.............. 500 

Petroleum refining ........................ ' 500 

Primary metals industries................. 200 

Food canning industries......... .......... 300 1 

Bottled and canned soft drinks............ 

* NAS, 1974 
water and make the iron unavailable to .plants. Such deficiencies 

induce chlorosis and further plant damage. Usually alkalinity 

must exceed 6 mg/L before such effects are noticed, however. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
*ALDRIN-DIELDRIN 

Aquatic Life 

Dieldrin 

For dieldrin the criterion to protect freshwater aquatic life 

as derived using the Guidelines is 0.0019 ug/L as a 24-hour 

average, and the concentration should not exceed 1.0 ug/L at any 

time. 

For dieldrin the criterion to protect saltwater aquatic life 

as derived using the Guidelines is 0.0019 ug/L as a 24-hour 

average, and the concentration should not exceed 0.71 ug/L at any 

time - 

Aldrin 

For freshwater aquatic life the concentration of aldrin 

should not exceed 4.0 ug/L at any time. No data are available 

concerning the chronic toxicity of aldrin to sensitive 

freshwater aquatic life. 

For saltwater aquatic life the concentration of aldrin should 

not exceed 1.3 ug/L at any time. No data are available 

concerning the chronic toxicity of aldrin to sensitive saltwater 

aquatic life. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to aldrin through ingestion of 

contaminated water and contaminated aquatic organisms, the 

I *Indicates suspended, canceled or restricted by U.S. EPA Office . 

of Pesticides and Toxic Substances 
a ambient water concentration should be zero, based on the 

nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, 

the levels which may result in incremental increase of cancer 

risk over the lifetime are estimated at and 

The corresponding recommended criteria are 0.74 ng/L, 0.074 ng/L, 

and 0.0074 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption 

of water, the levels are 0.79 ng/L, 0.079 ng/L, and 0.0079 

ng/L, respectively. 

FOP the maximum protection of human health from the potential 

carcinogenic effects of exposure to dieldrin through ingestion of 

contaminated water and Contaminated aquatic organisms, the 

ambient water concentration should be zero, based on the 

nonthreshold assumption for this chemical. However, zero level 

may not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

,lifetime are estimated at and The 

corresponding recommended criteria are 0.71 ng/L, 0.071 ng/L, and 

0.0071 ng/L, respectively. If these above estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 0.76 ng/L, 0.076 ng/L, and 0.0076 ng/L, 

respectively. 

0 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
AMMONIA 

SUMMARY : 

All concentrations used herein are expressed as un-ionized 

ammonia (NH3), because NH3, not the ammonium ion (NH4+) has been 

demonstrated to be the principal toxic form of ammonia. The 

data used in deriving criteria are predominantly from flow 

through tests in which ammonia concentrations were measured. 

Ammonia was reported to be acutely toxic to freshwater organisms 

at concentrations (uncorrected €or pH) ranging from 0.53 to 22.8 

mg/L NH3 for 19 invertebrate species representing 14 families and 

16 genera and from 0.083 to 4.60 mg/L NH3 €or 29 fish species 

from 9 families and 18 genera. Among fish species, reported 96- 

hour LC50 ranged from 0.083 to 1.09 mg/L for salmonids and from 

0.14 to 4.60 mg/L NH3 for nonsalmonids. Reported data from 

chronic tests on ammonia with two freshwater invertebrate 

species, both daphnids, showed effects at concentrations 

(uncorrected for pH) ranging from 0.304 to 1.2 mg/L NH3, and 

with nine freshwater fish species, from five families and seven 

genera, ranging from 0.0017 to 0.612 mg/L NH3. 

I 

Concentrations of ammonia acutely toxic to fishes may cause 

loss of equilibrium, hyperexcitability, increased breathing, 

cardiac output and oxygen uptake, and, in extreme cases, 

convulsions, coma, and death. At lower concentrations ammonia 

has many effects on fishes, including a reduction in hatching 

success, reduction in growth rate and morphological development, 

and pathologic changes in tissues of gills, livers, and kidneys. 
Several factors have been shown to modify acute NH3 toxicity 

in fresh water. Some factors alter the concentration of unionized ammonia in the water by affecting the aqueous ammonia 

equilibrium, and some factors affect the toxicity of un-ionized 

ammonia itself, either ameliorating or exacerbating the effects 

of ammonia. Factors that have been shown to affect ammonia 

toxicity include dissolved oxygen concentration, temperature, 

pH, previous acclimation to ammonia, fluctuating or intermittent 

exposures, carbon dioxide concentration, salinity, and the 

presence of other toxicants. 

The most well-studied of these is pH: the acute toxicity of 

NH3 has been shown to increase as pH decreases. Sufficient data 

exist from toxicity tests conducted at different pH values to 

formulate a mathematical expression to describe pH-dependent 

acute NH3 toxicity. The very limited amount of data regarding 

effects of pH on chronic NH3 toxicity also indicates increasing 

NH3 toxicity with decreasing pH, but the data are insufficient 

to derive a broadly applicable toxicity/pH relationship. Data on 

temperature effects on acute NH3 toxicity are limited and 

somewhat variable, but indications are that NH3 toxicity to fish 

is greater as temperature decreases. There is no information 

available regarding temperature effects on chronic NH3 toxicity. 

Examination of pH and temperature-corrected acute NH3 

toxicity values among species and genera of freshwater organisms 

showed that invertebrates are generally more tolerant than 

fishes, a notable exception being the fingernail clam. There is 

no clear trend among groups of fish; the several most sensitive 

a 
tested species and genera include representatives from diverse 

families (Salmonidae, Cyprinidae, Percidae, and Centrarchidae). 

Available chronic toxicity data for freshwater organisms also 

indicate invertebrates (cladocerans, one insect species) to be 

more tolerant than fishes, again with the exception of the 

fingernail clam. When corrected for the presumed effects of 

temperature andpH, there is also no clear trend among groups of 

fish for chronic toxicity values, the most sensitive species 

including representatives from five families (Salmonidae, 

Cyprinidae, Ictaluridae, Centrarchidae, and Catostomidae) and 

having chronic values ranging by not much more than a factor or 

two. The range of acute - chronic ratios for 10 species from 6 

0 

families was 3 to 43, and acute-khronic ratios were higher for 

the species having chronic tolerance below the median. 

Available data indicate that differences in sensitivities between 

warm and coldwater families of aquatic organisms are inadequate 

to warrant discrimination in the national ammonia criterion 

between bodies of water with I'warm1' and 'lcoldwaterll fishes; 

rather, effects of organism sensitivities on the criterion are 

most appropriately handled by site -specific criteria derivation 

procedures. 

Data for concentrations of NH3 toxic to freshwater 

phytoplankton and vascular plants, although limited, indicate 

that freshwater plant species are appreciably more tolerant to 

NH3 than are invertebrates or fishes. The ammonia criterion 

appropriate for the protection of aquatic animals will therefore 0 

in all likelihood be sufficiently protective of plant life. 

,, 
Available acute and chronic data for ammonia with saltwater 

organisms are very limited, and insufficient to derive a 

saltwater criterion. A few saltwater invertebrate species have 

been tested, and the prawn Macrobrachiurn rosenberqil was the 

most sensitive. The few saltwater fishes tested suggest greater 

sensitivity than freshwater fishes. Acute toxicity of NH3 

appears to be greater at low pH values, similar to findings in 

freshwater. Data for saltwater plant species are limited to 

diatoms, which appear to be more sensitive than the saltwater 

invertebrates for which data are available. 

More quantitative information needs to be published on the 

toxicity of ammonia to aquatic life. Several key research needs 

must be addressed to provide a more complete assessment of 

ammonia toxicity. These are: (1) acute tests with additional 

saltwater fish species and saltwater invertebrate species: (2) 

life-cycle and early life-stage tests with representative 

freshwater and saltwater organisms from different families, with 

particular attention to trends of acute-chronic ratios; (3) 

fluctuating and intermittent exposure tests with a variety of 

species and exposure patterns: (4) more complete tests of the 

individual and combined effects of pH and temperature, especially 

for chronic toxicity: (5) more histopathological and 

histochemical research with fishes, which would provide a rapid 

means of identifying and quantifying sublethal ammonia effects; 

and (6) studies on effects of dissolved and suspended solids on 

acute and chronic toxicity. 
NATIONAL CRITERIA: 

The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses indicate that, except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if: 

(1) the 1-hour* average concentration of un-ionized ammonia 

(in mg/L NH3) does not exceed, more often than once every 3 years 

on the average, the numerical value given by 0.52/FT/FPH/2, 

where: 

FT = 10-0.03(20-TCAP); TCAP - < T - < 30 

10-0.03(20-T) ; 0 - < T - < TCAP 

FPH = 1 :a<p~<g 

1+10-7.4-PH 

1.25 ; 6.5 

TCAP = 20 C: Salmonids or other 

coldwater species present 

= 25 C; Salmonids and other 

coldwater species absent 

<pH<8 

sensitive 

sensitive 

(*An averaging period of 1 hour may not be appropriate if 

excursions of concentrations to greater than 1.5 times the 

average occur during the hour; in such cases, a shorter averaging 

period may be needed.) 

(2) the 4-day average concentration of u n -ionized ammonia 

(in mg/L NH3) does not exceed, more 

on the average, the average* 

0.80/FT/FPH/RATIO, where FT and FPH , 

often than once every 3 years 

numerical value given by 

are as above and: 
RATIO = 16 ; 7.7 - < pH - <9 

= 24 10-7.7-ph 

1+10-7.4ph ;6.5< - ph - <7.7 

TCAP = 15 C; Salmonids or other sensitive 

coldwater species present 

coldwater species absent = 20 C: Salmonids and other sensitive 

(*Because these formulas are nonlinear in pH and temperature, the 

criterion should be the average of separate evaluations of the 

formulas reflective of the fluctuations of flow, pH, and 

temperature within the averaging period; it is not appropriate in 

general to simply apply the formula to average pH, temperature, 

and flow.) 

The extremes for temperature (0, 30) and pH (6.5, 9) given in 

the above formulas are absolute. It is not permissible with 

current data to conduct any extrapolations beyond these limits. 

In particular, there is reason to believe that appropriate 

criteria at pH > 9 will be lower than the plateau between pH 8 

and 9 given above. 

Criteria concentrations for the pH range 6.5 to 9.0 and the 

temperature range 0 C to 30 C are provided in the following 

tables. Total ammonia concentrations equivalent to each unionized ammonia concentration are also provided in these tables. 

There are limited data on the effect of temperature on chronic 

toxicity. EPA will be conducting additional research on the 

effects of temperature on ammonia toxicity in order to fill 

perceived data gaps. Because of this uncertainty, additional 

site-specific information should be developed before these 
criteria are used in wasteload allocation modeling. For example, 

the chronic criteria tabulated for sites lacking salmonids are 

less certain at temperatures much below 20 C than those tabulated 

at temperatures near 20 C. Where the treatment levels needed to 

meet these criteria below 20 C may be substantial, use of sitespecific criteria is strongly suggested. Development of such 

criteria should be based upon site-specific toxicity tests. 

Data available for saltwater species are insufficient to 

derive a criterion for saltwater. 

The recommended exceedence frequency of 3 years is the 

Agency's best scientific judgment of the average amount of time 

it will take an unstressed system to recover from a pollution 

event in which exposure to ammonia exceeds the criterion. A 

stressed system, for example, one in which several outfalls occur e 

in a limited area, would be expected to require more time for 

recovery. The resilience of ecosystems and their ability to 

recover differ greatly, however, and site-specific criteria may 

be established if adequate justification is provided. 

The use of criteria in designing waste treatment facilities 

requires the selection of an appropriate wasteload allocation 

model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

impractical, in which case one should rely on a steady-state 

model. The Agency recommends the interim use of 145 or lQlO for 

Criterion Maximum Concentration design flow and 745 or 7410 for 

the Criterion Continuous Concentration design flow in steady1 state models for unstressed and stressed systems respectively. 
(I) Onchow avuagm ancmntratlonr for mmonla.* 

,r\ 

PH oc 5c 10 c [IS c MC 25 c MC 

A. Salmmldr or omr Smnrltlvm Coldwatr Speclos Present 

Un-1onIz.d Amnia (mqllltor NH,) 

6.54 0.0091 0.0129 0.0182 0.026 O.OS6 0 .OY 0.036 

6.75 0.0149 0.02 I 0.030 0.042 0.059 0.059 O.OS9 

7 .oo 0.025 0.033 0 .Mb 0.066 0.09) 0.093 0.093 

0.054 0.048 0.068 0.095 0.135 0.135 0.135 

7 30 0.045 0 .w 0.091 0.128 0.181 0.181 0.181 

7 35 

n93 0.056 0.080 0.113 0.159 0.22 0.22 0.22 

8.00 0.065 0 092 0.130 0.184 0.26 0.26 0 36 

0.065 0.092 0.130 0.184 0.26 0.26 0.26 

0.065 0.092 0.130 0.184 0 -26 0 36 0.26 

8 35 

8.54 

8.15 0.065 0.092 0.130 0.184 0.26 0.26 0.26 

9 .oo 0.065 0.092 0.130 0.184 0 26 0 36 0 26 

Total Amonla (mg/Iltu NH3) 

35 

32 

28 

25 

17.4 

12.2 

8.0 

4 .I 

2.6 

I .47 

0 .a 

_- 33 

w) 

26 

22 

163 

11.4 

7 -5 

43 

2.4 

1 .w 

0.83 

31 

28 

25 

20 

15.5 

10.9 

7.1 

4.1 

2.3 

1.37 

0 A3 

30 

n -. 

24 

19.7 

14.9 

103 

6 9 

4 ;o 

2.3 

I .s 

0 .86 

29 

-. 27 

23 

19 .2 

14.6 

10.3 

6 . -0 

_. 

3.9 

2.3 

1.42 

0.91 

20 14.31 

18.6 13 3 

16.4 11.6 

13.4 9.5 

10.1 7.3 

4.8 3 .5 

2 .8 2.1 

1.71 I .28 

I -07 0.83 

0.72 0 38 

7.2md 5.2 

8. Sallnonlds and OthU S.nrltlv* Coldratw Speclor Absmt 

UWl0nlz.d AnnuxlIa (mg/Iltr NH,) 

0.0091 0.0129 0.0182 0.026 0 .036 0.051 o.os1 

0.0149 0.02 1 0.030 0.042 O.OS9 0.084 0.084 

0.023 0.033 0.046 0.066 0.093 0.131 0.131 

0.034 0.048 0.068 0.095 0.135 0.190 0.190 

0.045 0.064 0.091 0.128 0.181 0.26 0.26 

0.056 0.000 0.1 13 0.159 0.22 0.32 0.32 

0.065 0.092 0.130 0.184 0.26 0 57 0 57 

0.065 0.092 0.1SO 0.184 0.26 0.37 0.37 

6 -54 

6.75 

7 .oo 

7 25 

7.50 - 

7.75 

8.00 

8.25 

830 

8.75 

9 .00 

_-. 

0 ;065 0 ;092 Oil30 Oil84 0.26 0 ;37 0;37 

0.065 0.092 0.130 0.184 0.26 0.37 0.37 

0.06s 0 492 0.130 0.184 0 36 0.37 0.37 

Total Amnia (mg/llter NWg) 

35 

32 

28 

23 

17.4 

12.2 

8 .o 

4.s 

2.6 

1.47 

0.86 

33 

30 

26 

22 

16.3 

11.4 

7.5 

4.2 

2 -4 

I .40 

0.83 

31 

28 

25 

20 

15.5 

10.9 

7.1 

4.1 

2.3 

1.31 

0.83 

30 

27 

24 

19.7 

14.9 

10.5 

6.9 

4.0 

2.3 

1 .JB 

0.86 

29 

27 

23 

19.2 

14.6 

105 

6.8 

3.9 

2.3 

1.42 

0.91 

29 

26 

23 

19.0 

!4 .s 

10.2 

6.8 

4 .O 

2.4 

1.52 

1.01 

20 

18.6 

16.4 

13.5- 

4. 

2.9 

1 .el 

1.18 

0.82 
(2) 4-day average cmcentratlons for annonIa.* 

PH oc 5c 10 c 15 C 20 c is c MC 

A. Salmlds or Other Sensltlve ColdMater SDecles Present 

Un-Ionized Amonla (rng/lltr NH3) 

x 

6 30 0.0007 

6.75 0.0012 

7 .00 0.0021 

7.25 0.0037 

7 .so 0.0066 

7.75 0.0109 

8 .oo 0.0126 

8.25 0.0126 

830 0.0126 

8.75 0.0126 

9 -00 0.0126 

0.0009 

0.0017 

0.0029 

0.0052 

0.0093 

0.01 53 

0.0177 

0.0177 

0.0177 

0.0171 

0.0177 

0.0013 

0.0023 

0.0042 

0.0074 

0.0132 

0.022 

0.025 

0.025 

0.025 

0.025 

0.025 

0.0019 

0.0033 

0.0059 

0.0105 

0.0186 

0.03 1 

0.035 

0.035 

0.035 

0.035 

0.035 

0.0019 

0.0033 

0.0059 

0.0105 

0.01% 

0.031 

0.035 

0.035 

0.035 

0.035 

0 .Of5 

0.0019 0.0019 

0.0033 0.0033 

O.CO59 0.0059 

0.0105 0.0105 

0.0186 0.0186 

0.031 0.031 

0.035 0.035 

0.035 0.035 

0.035 0.035 

0.03s 0.035 

0.035 0.035 

6 30 23 2.4 2.2 23 1.49 1 .04 0.73 

6.73 2.5 2.4 2 .2 2.2 1.49 I .04 0.73 

7.00 2.5 2.4 23 23 1-49 1 .w 0.74 

7.25 2.5 2.4 2.2 2.2 1 .M 1 .04 0.74 

7.50 2.5 2.4 22 23 1 .M 1.05' 0.74 

7.75 2.3 2.2 2.1 2.0 1.40 0.99 0.71 

8 .a0 1.53 1.44 1.37 1.33 0.93 0.47 

8.25 0.87 0.82 0.78 0.76 0.54 

8.50 0.49 0.47 0.45 0.44 0 22 0 23 0.17 

8 .75 0.28 0.27 0.26 0.27 0.19 OS15 0.11 

9 .W 0.16 0.16 0.16 0.16 0.13 0.10 0.08 

00:s 0.28 

8. Salmonlds'and Other Sensltlve Coldwater Species Absent? 

Un-lonired'A~la (mg/lltr NH3) 

6.50 0.0007 0.0009 0.0013 0.0019 0.0026 0.0026 0.0026 

6.75 0.0012 0.0017 0.0023 0.0033 O.oQ47 0.0047 0-0047 

7 .00 0.0021 0.0029 0.0042 0.0059 0.0083 0.0083 0.0083 

7.25 0.0037 0.0052 0.0074 0.0105 0.0148 0.0148 0.0148 

7 30 0.0066 0.0093 0.0132 0.0186 0.026 0.026 0.026 

7.75 0.0109 0.01% 0.022 0.031 0.043 0.043 0.043 

8.00 0.0126 0.0177 0.025 0.035 0.050 0.050 0.050 

8.25 0.0126 0.0177 0.025 0.035 0.050 0.050 0.050 

8.50 0.0126 0.0177 0.025 0.035 0 .ow 0.0% 0.0% 

8.75 0.0126 0.0177 0.025 0.035 0.050 0.0% 0.oso 

9.W 0.0126 0.0177 0.025 0.035 0.0% 0.0% 0.0% 

6.50 

6.7s 

7 .00 

7 2s 

7 30 

7.75 

8 .oo 

8.25 

8.50 

8.75 

9 .oo 

2.5 

2.5 

2.4 

2.5 

2 5 

2.3 

1.53 

037 

0.49 

0.28 

0.16 

2 .4 

2 -4 

2 -4 

2 -4 

2.4 

2 -2 

1 -44 

0.82 

0.47 

0 37 

0.16 

2 .2 

2 3 

23 

2.2 

2.2 

2.1 

1.37. 

0.78 

0 -45 

0.26 

0.16 

23 

2 3 

23 

2-2 

23 

2.0 

1.33 

0.76 

0.44 

0.27 

0.16 

2.1 

2.1 

2.1 

2. I 

2.1 

1.98 

1.31 

0.76 

0.45 

037 

0.17 

Bs d 

1.46 l,ZD 1.03 0 8c 

1.47 1.z: 1.04 D.@g 

1.47 , tl 1 .04 0,BC 

1.48 l>2P 1.05 0 0(P 

1.491.22. I.WO,@? 

I.39 /,/U 1.00 0.82. 

0.93 0.7@ 061 Oa<s' 

034 0.40 

0 *33 0 35 

0.21 0.16 

0.14 0.1 I 

+ To convert ihw values to mg/llter N. multlply by 0.822. 

t Sltcspocltlc erltala dwelopmeni Is shongly suggested at tqm*Yros above 20 C 

because of ihe llmltd data aveilable to gwuate the crltrla rac-tlan, and 

at tanporatures b.lW 20 C boCauSe Of tha llnltsd data and brause WO aangn In 

the crltwla mq have algnltlcant Impact on the level of trmtunt rqulrd In 

meeting the recamended crltorie. 
The Agency acknowledges that the Criterion Continuous 

Concentration stream flow averaging period used for steady-state 

wasteload allocation modeling may be as long as 30 days in 

situations involving POTWs designed to remove ammonia where 

limited variability of effluent pollutant concentration and 

resultant concentrations in receiving waters can be demonstrated. 

In cases where low variability can be demonstrated, longer 

averaging periods for the ammonia Criterion Continuous 

Concentration (e.g., 30-day averaging periods) would be 

acceptable because the magnitude and duration of exceedences 

above the Criterion Continuous Concentration would be 

sufficiently limited. These matters are discussed in more detail 

in the Technical Support Document for Water Quality-Based Toxics 

Control (U.S. EPA, 1985a). 

(50 F.R. 30784, July 29, 1985) 

SEE APPENDIX A FOR METHODOLOGY 
ANTIMONY 

CRITERIA: 

Aquatic Life 

The available data for antimony indicate that acute and 

chronic toxicity to freshwater aquatic life occur at 

concentrations as low as 9,000 and 1,600 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. Toxicity to algae occurs at 

concentrations as low as 610 ug/L. 

No saltwater organisms have been adequately tested with 

antimony, and no statement can be made concerning acute or 

chronic toxicity. 

Human Health 

For the protection of human health from the toxic properties 

of antimony ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 146 

UWL. 

For the protection of human health from the toxic 

properties of antimony ingested through contaminated aquatic 

organisms alone, the ambient water criterion is determined to be 

45 mg/L. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 

J 
ARSENIC 0 AQUATIC LIFE SUMMARY: 

The chemistry of arsenic in water is complex and the form 

present in solution is dependent on such environmental conditions 

as Eh, pH, organic content, suspended solids, and sediment. The 

relative toxicities of the various forms of apsenic apparently 

vary from species to species. For inorganic arsenic(II1) acute 

values for 16 freshwater animal species ranged from 812 ug/L for 

a cladoceran to 97,000 ug/L for a midge, but the three acutechronic ratios only ranged from 4.660 to 4.862. The five acute 

values for inorganic arsenic(V) covered about the same range, but 

the single acute-chronic ratio was 28.71. The six acute values 

for MSMA ranged from 3,243 to 1',403,000 ug/L. The freshwater 

residue data indicated that arsenic is not bioconcentrated to a 

high degree but that lower forms of aquatic life may accumulate 

higher arsenic residues than fish. The low bioconcentration 

factor and short half-life of arsenic in fish tissue suggest that 

residues should not be a problem to predators of aquatic life. 

a 

The available data indicate that freshwater plants differ a 

great deal as to their sensitivity to arsenic(II1) and 

arsenic(V). In comparable tests, the alga, -----___--- Selenastrum 

capricornutum, was 45 times more sensitive to arsenic(V) than to 

arsenic(III), although other data present conflicting 

information on the sensitivity of this alga to arsenic(V). Many 

plant values for inorganic arsenic(II1) were in the same range as 

the available chronic values for freshwater animals; several 0 - 
plant values for arsenic(V) were lower than the one available 

chronic value. 

The other toxicological data revealed a wide range of 

toxicity based on tests with a variety of freshwater species and 

endpoints. Tests with early life stages appeared to be the most 

sensitive indicator of arsenic toxicity. Values obtained from 

this type of test with inorganic arsenic(II1) were lower than 

chronic values contained in Table 2. For example, an effect 

concentration of 40 ug/L was obtained in a test on inorganic 

arsenic(II1) with embryos and larvae of a toad. 

Twelve species of saltwater animals have acute values for 

inorganic arsenic(II1) from 232 to 16,030 ug/L and the single 

acute-chronic ratio is 1.945. The only values available for 

inorganic arsenic(V) are for two invertebrate and are between 

2,000 and 3,000 ug/L. Arsenic(II1) and arsenic(V) are equally 

toxic to various species of saltwater algae, but the 

sensitivities of the species range from 19 ug/L to more than 

1,000 ug/L. In a test with an oyster, a BCF of 350 was obtained 

for inorganic arsenic(II1). 

NATIONAL CRITERIA: 

The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses indicate that, except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the 4-day average concentration of arsenic(II1) 

does not exceed 190 ug/L more than once every 3 years on the 
average and if the 1-hour average concentration does not exceed 

360 ug/L more than once every 3 years on the average. 

The procedures described in the Guidelines indicate that, 

except possibly where a locally important species is very 

sensitive, saltwater aquatic organisms and their uses should not 

be affected unacceptably if the 4-day average concentration of 

arsenic(II1) does not exceed 36 ug/L more than once every 3 years 

on the average and if the 1-hour average concentration does not 

exceed 69 ug/L more than once every 3 years on the average. This 

criterion might be too high wherever Skeletonema cosrarum or 

Thalassiosira aestivalis --- are ecologically important. 

Not enough data are available to allow derivation of 

numerical national water quality criteria for freshwater aquatic 

life for inorganic arsenic(V) or any organic arsenic compound. 

Inorganic arsenic(V) is acutely toxic to freshwater aquatic 

animals at concentrations as low as 850 ug/L and an acute-chronic 

ratio of 28 was obtained with the fathead minnow. Arsenic(V) 

affected freshwater aquatic plants at concentrations as low as 48 

ug/L. Monosodium methanearsenace (MSMA) is acutely toxic to 

aquatic animals at concentrations as low as 1,900 ug/L, but no 

data are available concerning chronic toxicity to animals or 

toxicity to plants. 

Very few data are available concerning the toxicity of any 

form of arsenic other than inorganic arsenic(II1) to saltwater 

aquatic life. The available data do show that inorganic 

arsenic(v) is acutely toxic to saltwater animals at 

concentrations as low as 2,319 ug/L and affected some saltwater 
plants at 13 to 56 ug/L. 

chronic toxicity of any 

arsenic(II1) to saltwater 

No data are available concerning the 

form of arsenic other than inorganic 

aquatic life. 

EPA believes that a measurement such as tlacid-solublett would 

provide a more scientifically correct basis upon which to 

establish criteria for metals. The criteria were developed on 

this basis. However, at this time, no EPA approved methods for 

such a measurement are available to implement the criteria 

through the regulatory programs of the Agency and the States. 

The Agency is considering development and approval of methods for 

a measurement such as acid-soluble. Until available, however, 

EPA recommends applying the criteria using the total recoverable 

method. This has two impacts: (1) certain species of some metals 

cannot be analyzed directly because the total recoverable method 

does not distinguish between individual oxidation states, and (2) 

these criteria may be overly protective when based on the total 

recoverable method. 

The recommended exceedence frequency of 3 years is the 

Agency's best scientific judgment of the average amount of time 

it will take an unstressed system to recover from a pollution 

event in which exposure to arsenic(II1) exceeds the criterion. a 

stressed system, for example, one in which several outfalls occur 

in a limited area, would be expected to require more time for 

recovery. The resilience of ecosystems and their ability to 

recover differ greatly, however, and site -specific criteria may 

be established if adequate justification is provided. 

The use of criteria in designing waste treatment facilities 0 , 

requires the selection of an appropriate wasteload a1 location 
model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

impractical, in which case one should rely on a steady-state 

model, The Agency recommends the interim use of 145 or lQlO for 

Criterion Maximum Concentration design flow and 745 or 7Q10 for 

the Criterion Continuous Concentration design flow in steadystate models for unstressed and stressed systems respective1 y. 

These matters are discussed in more detail in the Technical 

Support Document fox Water Quality-Based Toxics Control (U.S. 

EPA, 1985). 

HUMAN HEALTH CRITERIA: 

For the maximum protection of human health from the potential 

carcinogenic effects due to expos,ure of arsenic through ingestion 

of contaminated water and contaminated aquatic organisms, the 

ambient water concentration should be zero based on the nonthreshold assumption for this chemical. However, zero level may 

not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

lifetime are estimated at and The corresponding 

criteria are 22 ng/L, 2.2 ng/L, and .22 ng/L, respectively. If 

the above estimates are made for consumption of aquatic 

organisms only, excluding consumption of water, the levels are 

175 ng/L, 17.5 ng/L, and 1.75 ng/L, respectively. Other 

concentrations representing different risk levels may be 

e 

calculated by use of the Guidelines. The risk estimate range is 

presented for information purpoes and does not represent an 

,, Agency judgment on an 81acceptable8t risk level. 
(45 

SEE 

F.R. 79318 

APPENDIX A 

Nov. 28,1980) (50 

FOR METHODOLOGY 

F.R. 30784, July 

a 
ASBESTOS 

CRITERIA: 

Aquatic Life 

No freshwater organisms have been tested with any asbestifom 

mineral and no statement can be made concerning acute or chronic 

toxicity 

No saltwater organisms have been tested with any asbestiform 

mineral and no statement can be made concerning acute or chronic 

toxicity. 

Human Health 

For the maximum protection of human health from the 

potential carcinogenic effects of exposure to asbestos through 

ingestion of water and contaminated aquatic organisms, the 

ambient water concentration should be zero. The estimated levels 

which would result in increased lifetime cancer risks of 

and are 300,000 fibers/L, 30,000 fibers/L, and 

3,000 fibers/L, respectively. Estimates for consumption of 

aquatic organisms only, excluding the consumption of water cannot 

be made. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR MmnoDomeY 
BACTERIA 

Freshwater Bathinq 

Based on a statistically sufficient number of samples 

(generally not less than 5 samples equally spaced over a 30-day 

period), the geometric mean of the indicated bacterial densities 

should not exceed one or the other of the following:(') 

E. coli 126 per 100 ml; or 

enterococci 33 per 100 ml; 

no sample should exceed a one sided confidence limit (C.L.) 

calculated using the following as guidance: 

designated bathing beach 75% C.L. 

moderate use for bathing 82% C.L. 

light use for bathing 90% C.L. 

infrequent use for bathing 95% C.L. 

based on a site-specific log standard deviation, or if site data 

are insufficient to establish a log standard deviation, then 

using 0.4 as the log standard deviation for both indicators. 

Marine Water Bathinq 

Based on a statistically sufficient number of samples 

(generally not less than 5 samples equally spaced over a 30-day 

period), the geometric mean of the enterococci densities should 

not exceed 35 per 100 ml; no sample should exceed a one sided 

confidence limit using the following as guidance: 

designated bathing beach 75% C.L. 

moderate use for bathing 82% C.L. 

light use for bathing 90% C.L. 

infrequent use for bathing 95% C.L. 
based on a site-specific log standard deviation, or if site data 

are insufficient to establish a log standard deviation, then 

using 0.7 as the log standard deviation. 

Note (1) - Only one indicator should be used. The Regulatory 

agency should select the appropriate indicator for its 

conditions. 

Shellfish Harvesting Waters 

The median fecal coliform bacterial concentration should not 

exceed 14 MPN per 100 ml with not more than 10 percent of samples 

exceeding 43 MPNperlOOml forthetakingof shellfish. 

RATIONALE 

Bathing Waters 

A recreational water quality criterion can be defined as a 

"quantifiable relationship between the density of an indicator in 

the water and the potential human health risks involved in the 

water's recreational use." From such a definition, a criterion 

can be adopted which establishes upper limits for densities of 

indicator bacteria in waters that are associated with acceptable 

health risks for swimmers. 

The Environmental Protection Agency, in 1972, initiated a 

series of studies at marine and fresh water bathing beaches which 

were designed to determine if swimming in sewage-contaminated 

marine and fresh water carries a health risk for bathers; and, if 

so, to what type of illness. Additionally, the Agency wanted to 

determine which bacterial indicator is best correlated to 

swimming-associated health effects and if the relationship is 

strong enough to provide a criterion. (l) 
The quantitative relationships between the rates of swimmingassociated health effects and bacterial indicator densities were 

determined using regression analysis. Linear relationships were 

estimated from data grouped on the basis of summers or trials 

with similar indicator densities. The data for each summer were 

analyzed by pairing the geometric mean indicator density for a 

summer bathing season at each beach with the corresponding 

swimming-associated gastrointestinal illness rate for the same 

summer. The swimming-associated illness rate was determined by 

subtracting the gastrointestinal illness rate in nonswimmers from 

that for swimmers. Ihese two variables from multiple beach sites 

were used to calculate a regression coefficient, y-intercept and 

95% confidence intervals for the paired data. In the marine 

studies the total number of points for use in regression analysis 

was increased by collecting trial days with similar indicator 

densities from each study location and placing them into groups. 

The swimming-associated illness rate was determined as before, by 

suk&racting the nonswimmer illness rate of all the individuals 

included in the grouped trial days from the swimmer illness rate 

during these safe grouped trial days. The grouping by trial days 

with similar indicator densities approach was not possible with 

the freshwater data because the variation of bacterial indicator 

densities in freshwater samples was not larqe enouqh to allow 

such an adjustment to be made. For the saltwater studies the 

results of the regression analyses of illness rates against 

indicator density data was very similar using the "by summer" or 0 Itby grouped trial days" approaches. 1 
The methods used to enumerate the bacterial indicator 

densities which showed the best relationship to swimmingassociated gastroenteritis rates were specifically developed €or 

the recreational water quality studies.' 

These membrane filter methods have successfully undergone 

precision and bias testing by the EPA Environmental Monitoring 

and Support Laboratory. (2) 

Several monitoring situations to assess bacterial quality are 

encountered by regulatory agencies. The situation needing the 

most rigorous monitoring is the designated swimming beach. Such 

areas are frequently lifeguard protected, provide parking and 

other public access and are heavily used by the public. Public 

beaches of this type were used by EPA in developing the 

relationship described in this document. 

Other recreational activities may involve bodies of water 

which are regulated by individual State water quality standards. 

These recreational resources may be natural wading ponds used by 

children or waters where incidential full body contact occurs 

because of water skiing or other similar activities. 

EPA's evaluation of the bacteriological data indicated that 

using the fecal coliform indicator group at the maximum geometric 

mean of 200perlOOm1, recommended inQualiwCriteria ___- forwater 

would cause an estimated 8 illness per 1,000 swimmers at fresh 

water beaches and 19 illness per 1,000 swimmers at marine 

beaches. These relationships are only approximate and are based 

on applying ratios of the geometric means of the various 

indicators from the EPA studies to the 200 per 100 ml fecal 
coliform criterion. However, these are EPA's best estimates of 

the accepted illness rates for areas which apply the EPA fecal 

coliform criterion. 

The E. coli and enterococci criteria presented in Table 1 

were developed using these currently accepted illness rates. The 

equations developed by D~four(~) and Cabelli(*) were used to 

calculate the geometric mean indicator densities corresponding to 

the accepted gastrointestinal illness rates. These densities are 

for steady state dry weather conditions. The beach is in noncompliance with the criteria if the geometric mean of several 

bacterial density samples exceeds the value listed in Table 1. 

Noncompliance is also signaled by an unacceptably high value 

for any single bacterial sample. The maximum acceptable 

bacterial density for a single sample is set higher than that for 

the geometric mean, in order to avoid necessary beach closings 

based on single samples. In deciding whether a beach should be 

left open, it is the long term geometric mean bacterial density 

that is of interest. Because of day-to-day fluctuations around 

this mean, a decision based on a single sample (or even several 

samples) may be erroneous, i.e., the sample may exceed the 

recommended mean criteria even though the long-term geometric 

mean is protective, or may fall below the maximum even if this 

mean is in the nonprotective range. 

To set the single sample maximum, it is necessary to specify 

the desired chance that the beach will be left open when the 

protection is adequate. This chance, or confidence level, was 

based on Agency judgment. For the simple decision rule 

considered here, a smaller confidence level corresponds to a more 

/ 
stringent (i.e. lower) single sample maximum. Conversely, a 

greater confidence level corresponds to less stringent (i.e. 

higher) maximum values. This technique reduces the chances of 

single samples inappropriately indicating violations of the 

recommended criteria. 

By using a control chart analogy (5) and the actual log 

standard deviations from the EPA studies, single sample maximum 

densities for various confidence levels were calculated. EPA 

then assigned qualitative use intensities to those confidence 

levels. A low confidence level (75%) was assigned to designated 

beach areas because a high degree of caution should be used to 

evaluate water quality for heavily used areas. Less intensively 

used areas would allow less restrictive single sample limits. 

Thus, 95% confidence might be appropriate for swimmable water in 

remote areas. Table 1 summarizes the results of these 

calculations. These single sample maximum levels should be 

recalculated for individual areas if significant differences in 

log standard deviations occur. 

The levels displayed in Table 1 depend not only on the 

assumed standard deviation of log densities, but also on the 

chosen level of acceptable risk. While this level was based on 

the historically accepted risk, it is still arbitrary insofar as 

the historical risk was itself arbitrary. 

Currently EPA is not recommending a change in the stringency 

of its bacterial criteria for recreational waters. Such a change 

does not appear warranted until more information based on greater 

experience with the new indicators can be accrued. EPA and the 
State Agencies can then evaluate the impacts of change in terms 

of beach closures and other restricted uses. 

Shellfish Harvesting Waters 

The microbiological criterion for shellfish water quality has 

been accepted by international agreement to be 70 total coliforms 

per 100 ml, using a median MPN, with no more than 10 percent of 

the values exceeding 230 total coliforms. No evidence of disease 

outbreak from consumption of raw shellfish which were grown in 

waters meeting this bacteriological criterion has been 

demonstrated. This standard has proven to be a practical limit 

when supported by sanitary surveys of the growing waters, 

acceptable quality in she1 lfish meats, and good epidemiological 

evidence. However, evidence from field investigations suggests 

that not all total coliform occurrences can be associated with 

fecal pollution. Thus, attention has been directed toward the 

adoption of the fecal coliform test to measure more accurately 

the occurrence and magnitude of fecal pollution in shellfishgrowing waters. 

A series of studies was initiated by the National Shellfish 

Sanitation Program and data relating the occurrence of total 

coliforms to numbers of fecal coliforms were compiled. The data 

showthata 70 coliform MPNperlOOmlatthe 5Othpercentilewas 

equivalent to a fecal coliform MPN of 14 per 1JO ml. The data, 

therefore, indicate that a median value for a fecal coliform 

standard is 15 and the 9Othpercentile shouldnot exceed 43 for a 0 5-tubeI 3-dilution method. 

EPA is currently (1986) co-sponsoring, with the National 
Oceanic and Atmospheric Administration, research into the 

application of the enterococci and E. coli indicators for 

assessing the quality of shellfish harvesting waters. The Food 

and Drug Administration is also reviewing the results of these 

studies. A change to the new indicators may be forthcoming if 

the studies show a correlation between gastrointestinal disease 

and the consumption of raw shellfish from waters with defined 

densities of the new indicators. However, these studies have not 

sufficiently progressed to justify any change at this time. 

Thus, evaluation of the microbiological suitability of waters for 

recreational taking of shellfish should be based upon the fecal 

colifonn bacterial levels. c6) 
CIIITERIA FOR INDICAllR FOR BrnIOrnCAL IEWITIES 

shle sanple Maximun Allowable Oensity (41, (5) 

Acceptable Srhh Steady State Designated Msderate Full Lightly Used Infrequmtly Used 

enteritis Fate per Indicator (upper 75% C.L.) Recreation Contact Rxreation 

1000 swimners Density (upper 82% CL.) Rereation (upper 95% CL.) 

Assaiated Qstro- Geawtric Mean madl Area my contact Full my m11 my ONltact 

(upr 90% C. L. 

Freshwater 

enterococci 8 33(1) 

-E. - coli 8 126P) 

Marine Bter 

61 

235 

89 108 151 

298 406 576 

enterococci 19 35(3) 104 124 276 

NOtest 

(1) Calculated to nearest mle nunber using quation: 

Calculated to nearest whole nunber using equation: 

Calculated to nearest wble nunber using equation: 

(mean entemxcci density) = i"tilogl0 

(mean -- E. coli density) = antFloglO 

(mean enterococci density) = antiloglo 

illness rate/lOOO people + 6.28 

9.40 

(2) 

illness rate/l000 pqi le + 11.74 

9.40 

(3) 

illness rate/lOOO people - 0.20 

12.17 

500 

(4) sirgle sanple lhit=mtilql0 (laglo indicator geamtric + Factor determined frm x (lq110 stand, 

mean density/100 ml) areas rnder the Wnnal deviation) 

prcbability cuwe for 

tb ass- level of 

pmbability 

The appmpdata factors for tb indicated one sided confidence levels are: 

75% C.L. - -675 

82% C.L. - -935 

90% C.L. - 1.28 

95% C.L. - 1.65 

(5) Based on the observed log standard deviations dwhq tk EPA studies: 

and enterccccci; and 0.7 for mq*e mter: enterococci. 

standard dwiation for its coditions &ich muld then vary t& sig(ie mple m t. 

0.4 for freshwater E. & 

Each 'uridi tion sbup fstablish-its om 
1. 

2. 

3. 

4. 

5. 

6. 

References 

Ambient Water Quality Criteria for Bacteria - 1986, EPA 

440/5-84-002, U.S. Environmental Protection Agency, Office 

of Water Regulations and Standards, Washington, DC. (NTIS 

access #: PB 86-158-045) 

Test Methods for Escherichia coli and Enterococii in Water 

By The Membrane Filter procedure, EPA 600/4-85-076, U.S. 

Environmental Protection Agency, Cincinnati, OH. (NTIS 

access #: PB 86-158-052) 

Dufour, A. P. 1983. Health Effects Criteria for Fresh 

Recreat iona 1 Waters. EPA- 6 0 O/ 1-8 4 -0 04 , U. S. Env ironmenta 1 

Protection Agency, Cincinnati, OH. 

Cabelli, V. J. 1981. Health Effects Criteria for Marine 

Recreational Waters. EPA-600/1-80-031, U.S. Environxental 

Protection Agency, Cincinnati, OH. 

ASTM. 1951. Manual on Quality Control of Materials. 

Special Technical Publication 15-C, American Society for 

Testing and Materials, Philadelphia, PA. 

U.S. Environmental Protection Agency. 1976. Quality 

Criteria for Water. U.S. Environmental Protection Agency, 

Washington, DC. 
BARIUM 

1 mg/L for domestic water supply (health). 

INTRODUCTION: 

Barium is a yellowish- white metal of the alkaline earth 

group. It occurs in nature chiefly as barite, BaS04 and 

witherite, BaC03, both of which are highly insoluble salts. The 

metal is stable in dry air, but readily oxidized by humid air or 

water. 

Many of the salts of barium are soluble in both water and 

acid, and soluble barium salts are reported to be poisonous 

(Lange, 1965: NAS, 1974). However, barium ions generally are 

thought to be rapidly precipitated or removed from solution by 

absorption and sedimentation (McKee and Wolf, 1963 NAS, 1974). 

I 

While barium is a malleable, ductile metal, its major 

commercial value is in its compounds. Barium compounds are used 

in a variety of industrial applications including the 

metallurgic, paint, glass and electronics industries, as well as 

for medicinal purposes. 

RATIONALE: 

Concentrations of barium drinking water supplies generally 

range from less than 0.6 ug/L to approximately 10 ug/L with upper 

limits ina fewmidwesternandwesternStates ranging from100to 

3,000 ug/L (PHS, 1962/1963; Katz, 1970; Little, 1971). Barium 

enters the body primarily through air and water, since 0 appreciable amounts are not contained in foods (NAS, 1974). 
The fatal dose of barium for man is reported to be 550 to 600 

mg. Ingestion of soluble barium compounds may also result in 

effects on the gastrointestinal tract, causing vomiting and 

diarrhea, and on the central nervous system, causing violent 

tonic and clonic spasms followed in some cases by paralysis 

(Browning, 1961; Patty, 1962, cited in Preliminary Air Pollution 

Survey of Barium and Its Compounds, 1969). Barium salts are 

considered to be muscle stimulants, especially for the heart 

muscle (Sollman, 1957). By constricting blood vessels, barium 

may cause an increase in blood pressure. On the other hand, it 

is not likely that barium accumulates in the bone, muscle, kidney 

or other tissues because it is readily excreted (Browning, 1961; 

McKee and Wolf, 1963). 

Stokinger and Woodward (1958) developed a safe concentration 

for barium in drinking water based on the limiting values for 

industrial atmospheres, an estimate of the amount absorbed into 

the blood stream, and daily consumption of 2 liters of water. 

From other factors they arrived at a limiting concentration of 2 

mg/L for a healthy adult human population, to which a safety 

factor was applied to allow for any possible accumulation in the 

body. Since barium is not removed by conventional water 

treatment processes and because of the toxic effect on the heart 

andbloodvessels,a limit oflmg/L is recommended for barium in 

domestic water supplies. 

Experimental data indicate that the soluble barium 

concentration in fresh and marine water generally would have to 

exceed 50 mg/L before toxicity to aquatic life would be expected. 

In most natural waters, there is sufficient sulfate or carbonate 
to precipitate the barium present in the water as a virtually 

insoluble, non-toxic compound. Recognizing that the physical and 

chemical properties of barium generally wil P preclude the 

existence of the toxic soluble form under usual marine and fresh 

water conditions, a restrictive criterion for aquatic life 

appears unwarranted. 

0 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
CRITERIA: 

Aquatic Life 

The available data for benzene indicate that acute toxicity 

to freshwater aquatic life occurs at concentrations as low as 

5,300 ug/L and would occur at lower concentrations among species 

that are more sensitive than those tested. No data are available 

concerning the chronic toxicity of benzene to sensitive 

freshwater aquatic life. 

The available data for benzene indicate that acute toxicity 

to saltwater aquatic life occurs at concentrations as low as 

5,100 ug/L and would occur at lower concentrations among species 

that are more sensitive than those tested. No definitive data 

are available concerning the chronic toxicity of benzene to 

sensitive saltwater aquatic life, but adverse effects occur at 

concentrations as lowas700ug/Lwitha fish species exposed for 

168 days. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to benzene through ingestion of 

contaminated water and contaminated aquatic organisms, the 

ambient water concentrations should be zero, based on the non 

threshold assumption Tor this chemical. However, zero level may 

not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

lifetime are estimated at and The 

corresponding recommended criteria are 6.6 ug/L, 0.66 ug/L, and 
0.066 ug/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of i 

water, the levels are 400 ug/L, 40.0 ug/L, and 4.0 ug/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
BENZIDINE 

Aquatic Life 

The available data for beniidine indicate that acute toxicity 

to freshwater aquatic life occurs at concentrations as low as 

2,500 ug/L and would occur at lower concentrations among species 

that are more sensitive than those tested NO data are available 

concerning the chronic toxicity of benzidine to sensitive 

freshwater aquatic life. 

No saltwater organisms have been tested with benzidine and no 

statement can be made concerning acute and chronic toxicity. 

Human Health 

For the maximum protection of human health from the 

potential carcinogenic effects of exposure to benzidine 

through ingestion of contaminated water and contaminated 

aquatic organisms, the ambient water concentrations should be 

zero, based on the nonthreshold assumption for this chemical. 

However, zero level may not be attainable at the present time. 

Therefore, the levels which may result in incremental increase 

of cancer risk over the lifetime are estimated at loq5, 

The corresponding recommended criteria are 

1.2 ng/L, 0.12 ng/L, and 0.01 ng/L, respectively. If these 

estimates are made for consumption of aquatic organisms only, 

excluding consumption of water, the levels are 5.3 ng/L, 0.53 

and 

ng/L, and 0.05 ng/L, respectively. 0 

145 F.R. 79318, November 28. 1980) _# 

SEE APPENDIX B ' FOR METHODOLOGY 
BERYLLIUM 

CRITERIA: 

Aquatic Life 

The available data for beryllium indicate that acute and 

chronic toxicity to freshwater aquatic life occur at 

concentrations as low as 130 and 5.3 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. Hardness has a substantial effect 

on acute toxicity. 

The limited saltwater data base available for beryllium does 

not permit any statement concerning acute or chronic toxicity. 

Human 

For the maximum protection of 

carcinogenic effects of exposure 

Health 

human health from the potential 

to beryllium through ingestion 

of contaminated water and contaminated aquatic organisms, the 

ambient water concentration should be zero, based on the non 

threshold assumption for this chemical. However, zero level may 

not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

lifetime are estimated at 10'-6t and 10-70 The 

corresponding recommended criteria are 37 ng/L, 3.7 ng/L, and 

0.37 ng/L, respectively. If these estimates are made for 
consumption of aquatic organisms only, excluding consumption of 

water, the levels are 641 ng/L, 64.1 ng/L, and 6.41 ng/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERION: 

750 mg/L for long-term irrigation on sensitive crops. 

INTRODUCTION: 

Boron is not found in its elemental form in nature: it is 

usually found as a sodium or calcium borate salt. Boron salts 

are used in fire retardants, the production of glass, leather 

tanning and finishing industries, cosmetics, photographic 

materials, metallurgy and for high energy rocket fuels. 

Elemental boron also can be used in nuclear reactors for neutron 

absorption. Borates are used as "burnable" poisons. 

RATIONALE : 

Boron is an essential element- for growth of plants but there 

is no evidence that it is required by animals, The maximum 

concentration found in 1,546 samples of river and lake waters 

from various parts of the United States was 5.0 mg/L; the mean 

Value was 0.1 mg/L (Kopp and Kroner, 1967). Ground waters could 

contain substantially higher concentrations at certain places. 

The concentration in seawater is reported as 4.5 mg/L in the form 

of borate (NAS, 1974). Naturally occurring concentrations of 

boron should have no effects on aquatic life. 

The minimum lethal dose for minnows exposed to boric acid at 

20 OC for 6 hours was reported to be 18,000 to 19,000 mg/L in 

distilled water and 19,000 to 19,500 mg/L in hard water (Le clerc 

and Devlaminck, 1955: Le Clerc, 1960). 
In the dairy cow, 16 to 20 g/day of boric acid for 40 days 

produced no ill effects (McKee and Wolf, 1963). 

Sensitive crops have shown toxic effects at 1000 ug/L or 

less of boron (Richards, 1954). Bradford (1966), in a review of 

boron deficiencies and toxicities, stated that when the boron 

concentration in irrigation waters was greater than 0.75 ug/L, 

some sensitive plants such as citrus began to show injury. 

Biggar and Fireman (1960) showed that with neutral and alkaline 

soils of high absorption capacities, water containing 2 ug/L 

boron might be used for some time without injury to sensitive 

plants. The criterion of 750 ug/L is thought to protect 

sensitive crops during long-term irrigation. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
CADMIUM 

AQUATIC - LIFE SUMMARY: 

Freshwater acute values for cadmium are available for species 

in 44 genera and range from 1.0 ug/L for rainbow trout to 28,000 

ug/L for a mayfly. The antagonistic effect of hardness on acute 

toxicity has been demonstrated with five species. Chronic tests 

have been conducted on cadmium with 12 freshwater fish species 

and 4 invertebrate species with chronic values ranging from 0.15 

ug/L for _- Daphnia -- mas5 - to 156 ug/L for the Atlantic salmon. 

Acute-chronic ratios are available for eight species and range 

from 0.9021 for the chinook salmon to 433.8 for the flagfish. 

Freshwater aquatic plants are affected by cadmium at 

concentrations ranging from 2 to 7,400 ug/L. These values are in 

the same range as the acute toxicity values for fish and 

invertebrate species, and are considerably above the chronic 

values. Bioconcentration factors (BCFs) for cadmium in fresh 

water range from 164 to 4,190 for invertebrates and from 3 to 

2,213 for fishes. 

Saltwater acute values for cadmium and five species of fishes 

range from 577 ug/L for larval Atlantic silverside to 114,000 

ug/L for juvenile mummichog. Acute values for 30 species of 

invertebrates range from 15.5 ug/L for a mysid to 135,000 ug/L 

for an oligochaete worm. The acute toxicity of cadmium 

generally increases as salinity decreases. The effect of 

temperature seems to be species-specif ic. Two life-cycle tests 

with - Mysidopsis bahia under different test conditions resulted in 

similar chronic values of 8.2 and 7.1 ug/L, but the acute-chronic 

ratios were 1.9 and 15, respectively. The acute values appear to 

0 
reflect effects of salinity and temperature, whereas the few 

available chronic values apparently do not. A life-cycle test 

with Mysidopsis bigelowi also resulted in a chronic value of 7.1 

ug/L and an acute-chronic ratio of 15. 

and macroalgae revealed effects at 22.8 to 860 ug/L. 

Studies with microalgae 

BCFs determined with a variety of saltwater invertebrates 

ranged from 5 to 3,160. BCFs for bivalve molluscs were above 

1,000 in long exposures, with no indication that steady-state 

had been reached. Cadmium mortality is cumulative for exposure 

periods beyond 4 days. Chronic cadmium exposure resulted in 

significant effects on the growth of bay scallops at 78 ug/L and 

on reproduction of a copepod at 44 ug/L. 

NATIONAL CRITERIA: 1 

The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses indicate that, except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the 4-day average concentration (in ug/L) of 

cadmium does not exceed the numerical value given by 

.(0.7852 [ ln(hardness) 1-3.490) more than once every 3 years on the 

average and if the one-hour average concentration (in ug/L) does 

not exceed the numerical value given by .(1.128[ ln(hardness)]- 

3.828) more than once every 3 years on the average. For 

example, at hardnesses of 50, 100, and 200 mg/L as CaC03 the 4- 

day average concentrations of cadmium are 0.66, 1.1, and 2.0 

ug/L, respectively, and the 1-hour average concentrations are 
1.8, 3.9 and 8.6 ug/L. If brook trout, brown trout, and striped 

bass are as sensitive as some data indicate, they might not be 

protected by this criterion. 

The procedures described in the Guidelines indicate that, 

except possibly where a locally important species is very 

sensitive, saltwater aquatic organisms and their uses should not 

be affected unacceptably if the 4-day average concentration of 

cadmium does not exceed 9.3 ug/L more than once every 3 years on 

the average and if the 1- hour average concentration does not 

exceed 43 ug/L more than once every 3 years on the average. The 

little information that is available concerning the sensitivity 

of the American lobster to cadmium indicates that this important 

species might not be protected by this criterion. In addition, 

data suggest that the acute toxicity of cadmium is salinity 

dependent: therefore, the 1-hour average concentration might be 

underprotective at low salinities and overprotective at high 

salinities. 

EPA believes that a measurement such as t'acid-solublell would 

provide a more scientifically correct basis upon which to 

establish criteria for metals. The criteria were developed on 

this basis. However, at this time, no EPA-approved methods for 

such a measurement are available to implement the criteria 

through the regulatory programs of the Agency and the States. 

The Agency is considering development and approval of methods for 

a measurement such as acid-soluble. Until available, however, 

EPA recommends applying the criteria using the total recoverable 0 

method. This has two impacts: (1) certain species of some metals 
cannot be analyzed directly because the total recoverable method 

does not distinguish between individual oxidation states, and (2) 

these criteria may be overly protective when based on the total 

recoverable method. 

The recommended exceedence frequency of 3 years is the 

Agency's best scientific judgment of the average amount of time 

it will take an unstressed system to recover from a pollution 

event in which exposure to cadmium exceeds the criterion. A 

stressed system, for example, one in which several outfalls occur 

in a limited area, would be expected to require more time for 

recovery. The resilience of ecosystems and their ability to 

recover differ greatly, however, and site-specific criteria may 

be established if adequate justification is provided. 

I 

The use of criteria in designing waste treatment facilities 

requires the selection of an appropriate wasteload allocation 

model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

impractical, in which case one should rely on a steady-state 

model. The Agency recommends the interim use of 145 or lQlO for 

Criterion Maximum Concentration design flow and 745 or 7410 for 

the Criterion Continuous Concentration design flow in steadystate models for unstressed and stressed systems, respectively. 

These matters are discussed in more detail in the Technical 

Support Document for Water Quality-Based Toxics Control (U.S. 

EPA, 1985). 

HUMAN HEALTH CRITERIA: 

The ambient water quality criterion for cadmium is 

recommended to be identical to the existing drinking water 
standard which is 10 ug/L. Analysis of the toxic effects data 0 

resulted in a calculated level which is protective of human 

health against the ingestion of contaminated water and 

contaminated aquatic organisms. The calculated value is 

comparable to the present standard. For this reason a selective 

criterion based on exposure solely from consumption of 6.5 grams 

of aquatic organisms was not derived. 

(45 F.R. 79318 Nov. 28,1980) (50 F.R. 30784, July 29, 1985) 

SEE APPENDIX A FOR METHODOLOGY 
CARBON TETRACHLORIDE 

CRITERIA: 

Aquatic Life 

The available data for carbon tetrachloride indicate that 

acute toxicity to freshwater aquatic life occurs at 

concentrations as low as 35,200 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity 

of carbon tetrachloride to sensitive freshwater aquatic life. 

The available data for carbon tetrachloride indicate that 

acute toxicity to saltwater aquatic life occurs at concentrations 

as low as 50,000 ug/L and would occur at lower concentrations 

among species that are more sensitive than those tested. No data 

are available concerning the chronic toxicity of 

carbontetrachloride to sensitive saltwater aquatic life. 

Human Health 

For the maximum protection of human health from the 

potential carcinogenic effects of exposure to carbon 

tetrachloride through ingestion of contaminated water and 

contaminated aquatic organisms, the ambient water 

concentrations should be zero, based on the nonthreshold 

assumption for this chemical. However, zero level may not be 

attainable at the present time. Therefore, the levels which may 

result in incremental increase of cancer risk over the 

lifetime are estimated at and lo-’. The 

corresponding recommended criteria are 4.0 ug/L, 0.40ug/L, and 

0.04 ug/L, respectively. If these estimates are made for 
consumption of aquatic organisms only, excluding consumption of 

I water, the levels are 69.4 ug/L, 6.94 ug/L, and 0.69 ug/L 

respectively. 

0 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
0 CRITERIA: 

CHLORDANE 

Aquatic Life 

For chlordane the criterion to protect freshwat tic life 

as derived using the Guidelines is 0.0043 ug/L as a 24-hOUr 

average, and the concentration should not exceed 2.4 ug/L at any 

time. 

For chlordane the criterion to protect saltwater aquatic life 

as derived using the Guidelines is 0.0040 ug/L as a 24-hour 

average, and the concentration should not exceed 0.09 ug/L at any 

time. 

Human Health 

r aqu 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to chlordane through ingestion 

of contaminated water and contaminated aquatic organisms, the 

ambient water concentration should be zero based on the 

nonthreshold assumption for this chemical. However, zero level 

may not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over 

the lifetime are estimated at loq5, lom6, and lo-’. The 

corresponding recommended criteria are 4.6 ng/L, 0.46 ng/L, and 

0.046 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 4.8 ng/L, 0.48 ng/L, and 0.048 ng/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CHLORINATED BENZENES 

CRITERIA: 

Aquatic - Life 

The available data for chlorinated benzenes indicate that 

acute toxicity to freshwater aquatic life occurs at 

concentrations as low as 250 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity of 

the more toxic of the chlorinated benzenes to sensitive 

freshwater aquatic life, but toxicity occurs at concentrations as 

low as 50 ug/L for a fish species exposed for 7.5 days. 

The available data for chlorinated benzenes indicate 

that acute and chronic toxicity to saltwater aquatic life occur 

at concentrations as low as 160 and 129 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

Human Health 

Xonochlorobenzene 

For comparison purposes, two approaches were used to derive 

criterion levels for monochlorobenzene. Based on available 

toxicity data, for the protection of public health the derived 

level is 488 ug/L. Using available organoleptic data, to 

control undesirable taste and odor quality of ambient water the 

estimated level is 20 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing 

water quality criteria, and have no demonstrated relationship to 

a potential adverse human health effects. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
0 CRITERIA: 

CHLQRINATED ETHANES 

Aquatic - Life 

The available freshwater data for chlorinated ethanes 

indicate that toxicity increases greatly with increasing 

chlorination, and that acute toxicity occurs at concentrations as 

low as 118,000 ug/L for l,Z-dichloroethane, 18,000 ug/L for two 

trichloroethanes, 9,320 ug/L for two tetrachloroethanes, 7,240 

ug/L for pentachloroethane, and 980 ug/L for hexachloroethane. 

Chronic toxicity occurs at concentrations as low as 20,000 ug/L 

for 1,2-dichloroethane, 9,400 ug/L for l,l,Z-trichloroethane, 

2,400 ug/L for 1,1,2,2-tetrachloroethane, 1,100 ug/L for 

pentachloroethane, and 540 ug/L for hexachloroethane. Acute and 

chronic toxicity would occur at lower concentrations among 0 species that are more sensitive than those tested. 

The available saltwater data for chlorinated ethanes indicate 

that toxicity increases greatly with increasing chlorination and 

that acute toxicity to fish and invertebrate species occurs at 

concentrations as low as 113,000 ug/L for lI2-dichloroethane, 

31,200 ug/L for l,l,l-trichloroethane, 9,020 ug/L for 

1,1,2,2-tetrachloroethane, 390 ug/L for pentachloroethane, 

and 940 ug/L for hexachloroethane. Chronic toxicity occurs at 

concentrations as low as 281 ug/L for pentachloroethane. Acute 

and chronic toxicity would occur at lower concentrations among 

species that are more sensitive than those tested. 
Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to 1,2-dichloroethane through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentration should be zero, based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at and The 

corresponding recommended criteria are 9.4 ug/L, 0.94 ug/L, and 

0.094 ug/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water,' the levels are 2,430 ug/L, 243 ug/L, and 24.3 ug/L, 

respectively. 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to l,l, 2-trichloroethane 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentration should be zero, based 

on the nonthreshold assumption for this chemical. However, 

zero level may not be attainable at the present time. Therefore, 

the levels which may result in incremental increase of cancer 

risk over the lifetime are estimated at and lo-'. 

The corresponding recommended criteria are 6.0 ug/L, 0.6 ug/L, 

and 0.06 ug/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 418 ug/L, 41.8 ug/L, and 4.18 ug/L, 

respectively. 
For the maximum protection of human health from the potential 

carcinogenic effects of exposure to 1,1,2,2-tetrachloroethane 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentration should be zero, based 

on the nonthreshold assumption for this chemical. However, 

zero level may not be attainable at the present time. Therefore, 

the levels which may result in incremental increase of cancer 

risk over the lifetime are estimated at and 

The corresponding recommended criteria are 1.7 ug/L, 0.17 ug/L, 

and 0.017 ug/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 107 ug/L, 10.7 ug/L, and 1.07 ug/L, 

respectively. I 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to hexachloroethane through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentration should be zero, based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at and loe7. The 

corresponding recommended criteria are 19 ug/L, 1.9 ug/L, and 

0.19 ug/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption 

of water, the levels are 87.4 ug/E, 8.74 ug/L, and 0.87 ug/L, 

respectively. 0 J 
For the protection of human health from the toxic properties 

of l,l,l-trichloroethane ingested through water and contaminated 

aquatic organisms, the ambient water criterion is determined to 

be 18.4 mg/L. 

For the protection of human health from the toxic properties 

of l,l,l-trichloroethane ingested through contaminated 

aquatic organisms alone, the ambient water criterion is 

determined to be 1.03 ug/l. 

Because of insufficient available data €or monochloroethane, 

1,l-dichloroethane, 1,1,l12-tetrachloroethane, and 

pentachloroethane, satisfactory criteria cannot be derived at 

this time, using the present guidelines. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERIA: 

CHLORINATED NAPHTHALENES 

Aquatic Life 

The available data for chlorinated naphthalenes indicate that 

acute toxicity to freshwater aquatic life occurs at 

concentrations as low as 1,600 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic 

toxicity of chlorinated naphthalenes to sensitive freshwater 

aquatic life. 

The available data for chlorinated naphthalenes, indicate that 

acute toxicity to saltwater aquatic life occurs at concentrations 

as low as 7.5 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of chlorinated 

naphthalenes to sensitive saltwater aquatic life. 

0 

Human Health 

Using the present guidelines, a satisfactory criterion cannot 

be derived at this time because of insufficient available data 

for chlorinated naphthalenes. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 0 2 
CHLORINE 

SUMMARY : 

Thirty -three freshwater species in 28 genera have been 

exposed to TRC and the acute values range from 28 ug/L for 

- Daenia -- magna to 710 Ug/L for the threespine stickleback. Fish 

and invertebrate species had similar ranges of sensitivity. 

Freshwater chronic tests have been conducted with two 

invertebrate and one fish species and the chronic values for 

these three species ranged from less than 3.4 to 26 ug/L, with 

acute- chronic ratios from 3.7 to greater than 78. 

The acute sensitivities of 24 species of saltwater animals in 

21 genera have been determined for CPO, and the LC50 range from 

26 ug/L for the eastern oyster to 1,418 ug/L for a mixture of two 

shore crab species. This range is very similar tO that observed 

with freshwater species, and fish and invertebrate species had 

similar sensitivities. Only one chronic test has been conducted 

with a saltwater species, Menidia Eeninsulae --I-__ I and in this test 

the acute chronic ratio was 1.162. 

The available data indicate that aquatic plants are more 

resistant to chlorine than fish and invertebrate species. 

NATIONAL CRITERIA: 

The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses indicate that, except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 
unacceptably if the 4-day average concentration of total residual 

chlorine does not exceed 11 ug/L more than once every 3 years on J-- 

the average and if the 1-hour average concentration does not 

exceed 19 ug/L more than once every 3 years on the average. 

The procedures described in the Guidelines indicate that, 

except possibly where a locally important species is very 

sensitive, saltwater aquatic organisms and their uses should not 

be affected unacceptably if the 4-day average concentration of 

chlorine - produced oxidants does not exceed 7.5 ug/L more than 

once every 3 years on the average and if the one-hour average 

concentration does not exceed 13 ug/L more than once every 3 

years on the average. 

The recommended exceedence frequency of 3 years is the 

Agency's best scientific judgment of the average amount of time 

, 

it will take an unstressed system to recover from a pollution 

event in which exposure to chlorine exceeds the criterion. A 

stressed system, for example, one in which several outfalls occur 

in a limited area, would be expected to require more time for 

recovery. The resilience of ecosystems and their ability to 

recover differ greatly, however, and site -specific criteria may 

be established if adequate justification is provided. 

The use of criteria in designing waste treatment facilities 

requires the selection of an appropriate wasteload a1 location 

model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

a impractical, in which case one should rely on a steady -state 

model. The Agency recommends the interim use of 145 or lQlO for 

Criterion Maximum Concentration design flow and 745 or 7Q10 for 
the Criterion Continuous Concentration design flow in steadystate models for unstressed and stressed systems, respectively. 

These matters are discussed in more detail in the Technical 

Support Document for Water Quality-Based Toxics Control (U.S. 

EPA, 1985). 

(50 F.R. 30784, July 29, 1985) 

SEE APPENDIX A FOR METHODOLOGY 
CHLORINATED PHENOLS 

Aquatic - Life 

The available freshwater data for chlorinated phenols 

indicate that toxicity generally increases with increasing 

chlorination, and that acute toxicity occurs at concentrations as 

low as 30 ug/L for 4-chloro-3-methylphenol to greater than 

500,000 ug/L for other compounds. Chronic toxicity occurs at 

concentrations as low as 970 ug/L for 2,4,6-trichlorophenol. 

Acute and chronic toxicity would occur at lower concentrations 

among species that are more sensitive than those tested. 

The available saltwater data for chlorinated phenols indicate 

that toxicity generally increases with increasing chlorination 

and that acute toxicity occurs at concentrations as low as 440 

ug/L for 2,3,5,6-tetrachlorophenol and 29,700 ug/L for 4- 

chlorophenol. Acute toxicity would occur at lower concentrations 

among species that are more sensitive than those tested. No data 

are available concerning the chronic toxicity of chlorinated 

phenols to sensitive saltwater aquatic life. 

Human Health 

Sufficient data are not available for 3-chlorophenol to 

derive a level which would protect against the potential toxicity 

of this compound. Using available organoleptic data, to 

control undesirable taste and odor qualities of ambient water, 

the estimated level is 0,1 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a a water quality criterion, and have no demonstrated / 
relationship to potential adverse human health effects. 

Sufficient data are not available for 4-chlorophenol to 

derive a level which would protect against the potential toxicity 

of this compound. Using available organoleptic data, to 

control undesirable taste and odor qualities of ambient water 

the estimated level is 0.1 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a 

water quality criterion, and have no demonstrated 

relationship to potential adverse human health effects. 

Sufficient data are not available for 2,3-dichlorophenol to 

derive a level which would protect against the potential toxicity 

of this compound. Using available organoleptic data, to control 

undesirable taste and odor qualities of ambient water the 

estimated level is 0.04 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a 

water quality criterion, and have no demonstrated relationship to 

potential adverse human health effects. 

Sufficient data are not available for 2,5-dichlorophenol to 

derive a level which would protect against the potential toxicity 

of this compound. Using available organoleptic data, to control 

undesirable taste and odor qualities of ambient water the 

estimated level is 0.5 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a 

water quality criterion, and have no demonstrated relationship to 

potential adverse human health effects. 

Sufficient data are not available for 2,6-dichlorophenol to 

derive a level which would protect against the potential toxicity 

of this compound. Using available organoleptic data, to control 
undesirable taste and odor qualities of ambient water the 

estimated level is 0.2 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a 

water quality criterion, and have no demonstrated relationship to 

potential adverse human health effects. 

Sufficient data are not available for 3,4-dichlorophenol to 

derive a level which would protect against the potential toxicity 

of this compound. Using available organoleptic data, to control 

undesirable taste and odor qualities of ambient water the 

estimated level is 0.3 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a 

water quality criterion, and have no demonstrated relationship to 

potential adverse human health ef'fects. 

For comparison purposes, two approaches were used to derive 

criterion levels for 2,4,5-trichlorophenol. Based on 

available toxicity data, to protect public health the derived 

level is 2.6 mg/L. Using available organoleptic data, to 

control undesirable taste and odor quality of ambient water the 

estimated level is 1.0 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a 

water quality criterion, and have no demonstrated relationship to 

potential adverse human health effects. 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to 2,4,6-trichlorophenoI 

through the ingestion of contaminated water and contaminated 

aquatic organisms, the ambient water concentration should be 

zero, based on the nonthreshold assumption for this 
chemical. However, zero level may not be attainable at the 

present time. Therefore, the levels which may result in 

incremental increase of cancer risk over the lifetime are 

estimated at 10-5, 10-6, and 10-7. The corresponding 

recommended criteria are 12 ug/L, 1.2 ug/L, and 0.12 ug/L, 

respectively. If these estimates are made for consumption of 

aquatic organisms only, excluding consumption of water, the 

levels are 36 ug/L, 3.6 ug/L, and 0.36 ug/L, respectively. Using 

available organoleptic data, to control undesirable taste and 

odor qualities of ambient water the estimated level is 2 ug/L. 

It should be recognized that organoleptic data have limitations 

as a basis for establishing. a water quality criterion, and 

have no demonstrated relationship to potential adverse 

human health effects. 

Sufficient data are not available for 2,3,4,6- 

tetrachlorophenol to derive a level which would protect against 

the potential toxicity of this compound. Using available 

organoleptic data, to control undesirable taste and odor 

qualities of ambient water the estimated level is 1.0 ug/L. It 

should be recognized that organoleptic data have limitations as a 

basis for establishing a water quality criterion, and have 

demonstrated relationship to potential adverse human health 

effects. 

Sufficient data are not available for 2-methyl-4-chlorophenol 

to derive a criterion level which would protect against any 

potential toxicity of this compound. Using available 

organoleptic data, to control undesirable taste and odor 

qualities of ambient water the estimated level is 1,800 ug/L. It 
should be recognized that organoleptic data have limitations as a 

basis for establishing a water quality criterion and have no 

demonstrated relationship to potential adverse human health 

effects . 

Sufficient data are not available for 3-methyl-4-chlorophenol 

to derive a criterion level which would protect against any 

potential toxicity of this compound. Using available 

organoleptic data, to control undesirable taste and odor 

qualities of ambient water the estimated level is 3,000 ug/L. It 

should be recognized that organoleptic data have limitations as a 

basis for establishing a water q u a l i t y c r i t e r i o n , and have no 

demonstrated relationship to potential adverse human health 

effects , 

Sufficient data are not available for 3-methyl-6-chlorophenol 

to derive a criterion level which would protect against any 

potential toxicity of this compound. Using available 

organoleptic data, to control undesirable taste and odor 

qualities of ambient water the estimated level is 20 ug/L. It 

should be recognized that organoleptic data have limitations as a 

basis for establishing a water quality criterion, and have no 

demonstrated relationship to potential adverse human health 

effects . 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
a CRITERIA : 

CHLOROALKYL ETHERS 

Aquatic - Life 

The available data for chloroalkyl ethers indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 238,000 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. NO 

definitive data are 

chloroalkyl ethers to sensitive freshwater aquatic life. 

available concerning the chronic toxicity of 

No saltwater organism has been tested with any chloroalkyl 

ether and therefore, no statement can be made concerning acute or 

chronic toxicity. 

Human Health 

For the protection of human health from the toxic properties 

of bis(2-chloroisopropy1) ether ingested through water and 

contaminated aquatic organisms, the ambient water criterion is 

determined to be 34.7 ug/L. 

For the protection of human health from the toxic properties 

of bis(2-chloroisopropy1) ether ingested through contaminated 

aquatic organisms alone, the ambient water criterion is 

determined to be 4.36 mg/L. 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to bis(chloromethy1) ether 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, based 

on the nonthreshold assumption for this chemical. However, zero 
level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at 10-5, and 

The corresponding recommended criteria are 37.6 x 

ug/L, 3.76 x ug/L, and 0.376 x ug/L, 

respectively. If these estimates are made for consumption of 

aquatic organisms only, excluding consumption of water, the 

levels are 18.4 x ug/L, 1.84 x ug/L, and 0.184 x 

10-3 ug/L, respectively. 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to bis(2-chloroethyl) ether 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at and 10-7. 

The corresponding recommended criteria are 0.30 ug/L, 0.030 

ug/L, and 0.003 ug/L, respectively. If these estimates are 

made for consumption of aquatic organisms only, excluding 

consumption of water, the levels are 13.6 ug/L, 1.36 ug/L, and 

0.136 ug/L, respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERIA: 

CHLOROFORM 

Aquatic Life 

The available data for chloroform indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 28,900 ug/L, and would occur at lower concentrations among 

species that are more sensitive than the three tested species 

Twenty-seven-day LC50 values indicate that chronic toxicity 

occurs at concentrations as low as 1,240 ug/L, and could occur at 

lower concentrations among species or other life stages that are 

more sensitive than the earliest life cycle stages of the rainbow 

trout. The data base for saltwater species is limited to one 

test and therefore, no statement can be made concerning acute or 

1 

chronic toxicity. 

Human Health 

For the maximum protection of human health from the 

potential carcinogenic effects of exposure to chloroform 

through ingestion of contaminated water and contaminated 

aquatic organisms, the ambient water concentrations should be 

zero, based on the nonthreshold assumption for this chemical. 

However, zero level may not be attainable at the present time. 

Therefore, the levels which may result in incremental increase 

of cancer risk over the lifetime are estimated at loe5, 

The corresponding recommended criteria are 

. 1.90 ug/L, 0.19 ug/L, and 0.019 ug/L, respectively. If these 

and 
estimates are made for consumption of aquatic organisms only, 

excluding consumption of water, the levels are 157 ug/L, 15.7 

ug/L, and 1.57 ug/L, respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CHMROPHENOXY HERBICIDES 

2,4-D; 2,4,5-TP 

CRITERIA: 

2,4-D 100 ug/L for domestic water supply (health) 

2,4,5-TP 10 ug/L for domestic water supply (health) 

RATIONALE: 

Two widely used herbicides are 2,4-D (2, 4- 

dichlorophenoxyacetic acid) and 2,4,5-TP (silvex) [2-(2’4, 5- 

trichlorophenoxy) propionic acid. Each of these compounds is 

formulated in a variety of salts and esters that may have a 

marked difference in herbicidal properties, but all are 

hydrolyzed rapidly to the corresponding acid in the body. 

The subacute oral toxicity of chlorophenoxy herbicides has 

been investigated in a number of species of experimental animals 

(Palmer and Radeleff, 1964; Lehman, 1965). The dog was found to 

be sensitive and often displayed mild injury in response to doses 

of 10 mg/kg/day for 90 days, and serious effects from a dose of 

20 mg/kg/day for 90 days. Lehman (1965) reported that the noeffect level of 2,4-D is 0.5 mg/kg/day in the rat, and 8.0 

mg/kg/day in the dog. 

Data are available on the toxicity of 2,4-D to man. A daily 

dosage of 500 mg (about 7 mg/kg) produced no apparent ill effects 

in a volunteer over a 21-day period (Kraus, 1946). When 2,4-D 

was investigated as a possible treatment €or disseminated 

coccidioidomycosis, the patient had no side effects from 18 

intravenous doses during 33 days; each of the last 12 doses in 

a 
the series was 800 mg (about 15 mg/kg) or more, the last being 

2000 mg (about 37 mg/kg) (Seabury, 1963). A 19th and final dose 

of 3600 mg (67 mg/kg) produced mild symptoms. 

The long-term no - effects levels (mg/kg/day) are listed for 

the rat and the dog. Those values are adjusted by a factor of 

1/500 for 2,4-D and 2,4,5-TP. The safe levels are then 

readjusted to reflect total allowable intake per person. Since 

little 2,4-D or 2,4,5-TP is expected to occur in foods, 20 

percent of the safe exposure level can reasonably be allocated to 

water without jeopardizing the health of the consumer. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
CHROMIUM rvI) 0 AQUATIC LIFE SUMMARY: 

Acute toxicity values for chromium(V1) are available for 

freshwater animal species in 27 genera and range from 23.07 

ug/L for a cladoceran to 1,870,000 ug/L for a stonefly. These 

species include a wide variety of animals that perform a wide 

spectrum of ecological functions. All five tested species of 

daphnids are especially sensitive. The few data that are 

available indicate that the acute toxicity of chromium(V1) 

decreases as hardness and pH increase. 

The chronic value for both rainbow trout and brook trout is 

264.6 uq/L, which is much lower than the chronic value of 1,987 

ug/L for the fathead minnow. The acute-chronic ratios for these 

three fishes range from 18.55 to 260.8. In all three chronic 

tests a temporary reduction in growth occurred at low 

concentrations. Six chronic tests with five species of daphnids 

gave chronic values that range from <2.5 to 40 ug/L and the 

acute-chronic ratios range from 1.130 to >9.680. Except for the 

fathead minnow, all the chronic tests were conducted in soft 

water. Green algae are quite sensitive to chromium(V1). The 

bioconcentration factor obtained with rainbow trout is less than 

3. Growth of chinook salmon was reduced at a measured 

concentration of 16 ug/L. 

0 

The acute toxicity of chromium (VI) to 23 saltwater vertebrate 

and invertebrate species ranges from 2,000 ug/L €or a polychaete 

worm and a mysid to 105,000 ug/L for the mud Snail. The chronic 

values €or a polychaete range from <13 to 36.74 ug/L, whereas 
that for a mysid is 132 ug/L. The acute-chronic ratios range 

from 15.38 to >238.5. Toxicity to macroalgae was reported at 

1,000 and 5,000 ug/L. Bioconcentration factors for chromium(V1) 

range from 125 to 236 for bivalve molluscs and polychaetes. 

CHROMIUM 1111) 

Acute values for chromium(II1) are available for 20 freshwater 

animal species in la genera ranging from 2,221 ug/L for a mayfly 

to 71,060 ug/L for caddisfly. Hardness has a significant 

influence on toxicity, with chromium(II1) being more toxic in 

soft water. 

A life-cycle test with __ Daphnia ____ __ maqna __ in soft water gave a 

chronic value of 66 ug/L. In a comparable test in hard water the 

lowest test concentration of 44 ug/L inhibited reproduction of 

- Dap&nia --- mgE, but this effect may have resulted from ingested 

precipitated chromium. In a life-cycle test with the fathead 

minnow in hard water the chronic value was 1,025 ug/L. Toxicity 

data are available for only two freshwater plant species. A 

concentration of 9,900 ug/L inhibited growth of roots of Eurasian 

watermilfoil. A freshwater green alga was affected by a 

concentration of 397 ug/L in soft water. No bioconcentration 

factor has been measured for chromium(II1) with freshwater 

organisms. 

Only two acute values are available for chromium (111) in 

saltwater 10,300 ug/L for the eastern oyster and 31,500 ug/L for 

the mummichog. In a chronic test effects were not observed on a 

polychaete worm at 50,400 ug/L at pH = 7.9, but acute lethality 

occurred when pH = 4.5. Bioconcentration factors for saltwater 
organisms and chromium(II1) range from 86 to 153, similar to the 

bioconcentration factors for chromium(V1) and saltwater species. 

NATIONAL CRITERIA: 

CHROMIUM (VI) 

The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses indicate that, except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the 4-day average concentration of chromium(V1) 

does not exceed 11 ug/L more than once every 3 years on the 

average and if the 1-hour avera e concentration does not exceed 7 

16 ug/L more than once every 3 years on the average. 

The procedures described in the Guidelines indicate that, 

except possibly where a locally important species is very 

sensitive, saltwater aquatic organisms, and their uses should not 

be affected unacceptably if the 4-day average concentration of 

chromium(V1) does not exceed 50 ug/L more than once every 3 years 

on the average and if the 1-hour average concentration does not 

exceed 1,100 ug/L more than once every 3 years on the average. 

Data suggest that the acute toxicity of chromium (VI) is salinity 

dependent; therefore, the 1-hour average concentration might be 

underprotective at low salinities. 

CHROMIUM( 1111 

The procedures described in the Guidelines indicate that, 

.I 

except possibly where a locally important species is very 

sensitive, freshwater aquatic organisms and their uses should not 
be affected unacceptably if the 4-day average concentration (in 

ug/L) of chromium(II1) does not exceed the numerical value given 

by e(0.8190[ln(hardness)]+l.561) more than once every 3 years on 

the average and if the 1-hour average concentration (in ug/L) 

does not exceed the numerical value given by 

(0.8190[ln(hardness) ]+3.688) more than once every 3 years on the e 

average. For example, at hardnesses of 50, 100, and 200 mg/L as 

CaC03 the &day average concentrations of chromium(II1) are 120, 

210, and 370 ug/L, respectively, and the 1-hour average 

concentrations are 980, 1,700, and 3,100 ug/L. 

No saltwater criterion can be derived for chromium(III), but 

10,300 ug/L is the EC50 for eastern oyster embryos, whereas 

50,400 ug/L did 

test. 

EPA believes 

provide a more 

not affect a polychaete worm in a life-cycle 

a 

that a measurement such as 18acid-soluble1' would 

scientifically correct basis upon which to 

establish criteria for minerals. The criteria were developed on 

this basis. However, at this time, no EPA-approved methods for 

such a measurement are available to implement the criteria 

through the regulatory programs of the Agency and the States. 

The Agency is considering development and approval of methods 

for a measurement such as acid-soluble. Until available, 

however, EPA recommends applying the criteria using the total 

recoverable method. This has two impacts: (1) certain species of 

some metals 

recoverable 

cannot be analyzed directly 

method does not distinguish 

because the total 

a between individual 
oxidation states, and (2) these criteria may 0 when based on the total recoverable method. 

The recommended exceedence frequency 

be overly protective 

of 3 years is the 

'Agency's best scientific judgment of the average amount of time 

it will take an unstressed system to recover from a pollution 

event in which exposure to chromium exceeds the criterion. A 

stressed system, for example, one in which several outfalls occur 

in a limited area, would be expected or require more time for 

recovery. The resilience of ecosystems and their ability to 

recover differ greatly, however, and site-specific criteria may 

be established if adequate justification is providedThe use of criteria in designing waste treatment facilities 

requires the selection of an appropriate wasteload a1 location 

model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

impractical, in which case one should rely on a steady-state 

model. The Agency recommends the interim use of 1Q5 or lQlO for 

Criterion Maximum Concentration design flow and 745 or 7410 or 

the Criterion Continuous Concentration design flow in steadystate models for unstressed and stressed systems, respectively. 

These matters are discussed in more detail in the Technical 

Support Document for water Quality-Based Toxics Control (U.S. 

EPA, 1985). 

HUMAN HEALTH CRITERIA: 

0 

For the protection of human health 

of Chromium I11 ingested through water 0 

from the toxic properties 

and contaminated aquatic 
organisms, the ambient water criterion is determined to be 170 

mg/ LJ. 

For the protection of human health from the toxic properties 

of Chromium I11 ingested through contaminated aquatic organisms 

alone, the ambient water criterion is determined to be 3433 mg/L. 

The ambient water quality criterion for total Chromium VI is 

recommended to be identical to the existing drinking water 

standard which is 50 ug/L. Analysis of the toxic effects data 

resulted in a calculated level which is protective of human 

health against the ingestion of contaminated water and 

contaminated aquatic organisms. The calculated value is 

comparable to the present standard. For this reason a selective 

criterion based on exposure solely from consumption of 6.5 grams 

of aquatic organisms was not derived. 

(45 F.R. 79318 Nov. 28,1980) (50 F.R. 30784, July 29, 1985) 

SEE APPENDIX A FOR METHODOLOGY 
2-CHLOROPHENOL 

Aquatic Life 

The available data for 2-chlorophenol indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 4,380 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No definitive 

data are available concerning the chronic toxicity of 2- 

chlorophenol to sensitive freshwater aquatic life, but flavor 

impairment occurs in one species of fish at concentrations as low 

as 2,000 ug/L. 

No saltwater organisms have been tested with 2- 

I 

chlorophenol and therefore, no statement can be made concerning 

acute or chronic toxicity. 

Human Health 

Sufficient data are not available for 2-chlorophenol to 

derive a level which would protect against the potential toxicity 

of this compound. Using avai 1 able organol eptic data, to 

control undesirable taste and odor qualities of ambient water the 

estimated level is 0.1 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a 

water quality criterion, and have no demonstrated relationship to 

potential adverse human health effects. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
COLOR 

Waters shall be virtually free from substances producing 

objectionable color for aesthetic purposes; 

the source of supply should not exceed 75 color units 

on the platinum-cobalt scale for domestic water 

supplies; and 

not reduce the depth of the compensation point for 

photosynthetic activity by more than 10 percent from 

the seasonally established norm for aquatic life. 

increased color (in combination with turbidity) should 

INTRODUCTION: 

Color in water principally results from degradation processes 

in the natural environment. Although colloidal forms of iron and 

manganese occasionally are the cause of color in water, the most 

common causes are complex organic compounds originating from the 

decomposition of naturally occurring organic matter (AWWA, 

1971). Sources of organic material include human materials from 

the soil such as tannins, human acid and humates: decaying 

plankton: and other decaying aquatic plants. Industrial 

discharges may contribute similar compounds: for example, those 

from the pulp and paper and tanning industries. Other industrial 

discharges may contain brightly colored substances such as those 

from certain processes in textile and chemical industries. 

Surface waters may appear colored because of suspended matter 

which comprises turbidity. Such color is referred to as apparent 

color and is differentiated from true color caused by colloidal 

human materials (Sawyer, 1960). Natural color is reported in , 
color 9mits" which generally are determined by use of the 

platinum-cobalt method (Standard Methods, 1971). 

There is no general agreement as to the chemical composition 

of natural color, and in fact the composition may vary chemically 

from place to place (AWWA, 1971). Black and Christman (1963a) 

characterized color-causing colloids examined as aromatic, 

polyhydroxy, methoxy carboxyl ic acids. Shapiro (1964) 

characterized color-causing constituents as being dialyzable and 

composed of aliphatic, polyhydroxyl carboxyl ic acids with 

molecular weights varying from less than 200 to approximately 

400. The colloidal fraction of color exists in the 3.5 to 10 mu 

diameter range (Black and Christman, 1963b). These same authors 

summarized other characteristics of color observed in laboratory 

studies of natural waters: color is caused by light scattering 

and fluorescence rather than absorption of light energy, and pH 

affects both particle size of the color-causing colloids and the 

intensity of color itself. 

RATIONALE : 

Color in water is an important constituent in terms of 

aesthetic considerations. To be aesthetically pleasing, water 

should be virtually free from substances introduced by man's 

activities which produce objectionable color. "Objectionable 

color18 is defined to be a significant increase over natural 

background levels. Non-natural colors such as dyes should not be 

perceptible by the human eye as such colors are especially 

objectionable to those who receive pleasure by viewing water in 

its natural state. Because of the extreme variations in the 
natural background amount of color, it is meaningless to attempt 

numerical limits. The aesthetic attributes of water depend on 

one’s appreciation of the water setting. 

0 

The effects of col’or on public water supplies also are 

principally aesthetic. The 1962 Drinking Water Standards (PHS, 

1962) recommended that color in finished waters should not exceed 

15 units on the platinum-cobalt scale. Water consistently can be 

treated using standard coagulation, sedimentation and filtration 

processes to reduce color to substantially less than 15 color 

units when the source water does not exceed 75 color units AWWA, 

1971; NAS, 1974). 

The effects of color in water on aquatic life principally are 

to reduce light penetration and thereby generally reduce 

photosynthesis by phytoplankton and to restrict the zone for a aquatic vascular plant growth. 

The light supply necessary to support plant life is dependent 

on both intensity and effective wave lengths (Welch, 1952). In 

general, the rate of photosynthesis increases with the intensity 

of the incident light. Photosynthetic rates are most affected in 

the red region and least affected in the blue-violet region of 

incident light (Welch, 1952). It has been found that in colored 

waters the red spectrum is not a region of high absorption so 

that the effective penetration, and therefore the intensity for 

photosynthesis, is not as restricted as are other wave lengths. 

It should be emphasized that transmission of all parts of the 

spectrum is affected by color, but the greatest effect is on the 0 standard or blue end of the spectrum (Birge and Juday, 1930). In 
TABLE 2. 

Maximum color of surface waters that have been 

used as sources for industrial water supplies. 

Industry or Industrial Use Color units 

Boiler make up 1,200 

cooling water 1,200 

Pulp and paper 

Chemical and allied products 

Petroleum 

360 

500 

25 
highly colored waters (45 to 132 color units) Birge and Juday 

(1930) measured the light transmission as a percentage of the 

incident level and found very little blue, 50 percent or less 

yellow, and 100 to 120 percent red. 

0 

The light intensity required for some aquatic vascular 

plants to photosynthetically balance the oxygen used in 

respiration may be 5 percent of full sunlight during maximum 

summer illumination periods (NTAC, 1968). As much as 10 percent 

of the incident light may be required for plankton to likewise 

photosynthetically produce sufficient oxygen to balance their 

respiration requirements (NTAC, 1968). The depth at which such a 

compensation point is reached, calledthe compensation depth, 

delineates the zone of ef fe,ctive photosynthetic oxygen 

production. To maintain satisfactory biological conditions, this 0 depth cannot be substantially reduced. 

Industrial requirements as related to water color have been 

standardized (NAS, 1974). Table 2 lists the maximum value used 

as a source of water for various industries and industrial uses. 

Through treatment, such waters can be made to meet almost any 

industrial requirement. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
0 AQUATIC LIFE SUMMARY: 

*COPPER 

Acute toxicity data 

freshwater animals. At 

are available for species in 41 genera of 

a hardness of 50 mg/L the genera range in 

sensitivity from 16.74 ug/L for Ptychocheilus - to 10,240 ug/L for 

Acroneuria. Data for eight species indicate that acute toxicity 

decreases as hardness increases. Additional data for several 

species indicate that toxicity also decreases with increases in 

alkalinity and total organic carbon. 

Chronic values are available €or 15 freshwater species and 

range from 3.873 ug/L €or brook trout to 60.36 ug/+ for northern 

pike. Fish and invertebrate species seem to be about equally 

sensitive to the chronic toxicity of copper. 

Toxicity tests have been conducted on copper with a wide 

range of freshwater plants and the sensitivities are similar to 

those of animals. Complexing effects of the test media and a 

lack of good analytical data make interpretation and application 

of these results difficult. Protection of animal species, 

however, appears to offer adequate protection of plants. Copper 

does not appear to bioconcentrate very much in the edible portion 

of freshwater aquatic species. 

The acute sensitivities of saltwater animals to copper range 

from 5.8 ug/L for the blue mussel to 600 ug/L for the green crab. 

A chronic life-cycle test has been conducted with a mysid, and 

adverse effects were observed at 77 ug/L but not at 38 ug/L, 

which resulted in an acute-chronic ratio of 3.346. Several 

*Indicates susDended. canceled or restricted bv U.S.EPA Office 

of Pesticides ahd Toxic Substances 

- 
saltwater algal species have been tested, and effects were 

observed between 5 and 100 ug/L. Oysters can bioaccumulate 

copper up to 28,200 times, and become bluish-green, apparently 

without significant mortality. In long-term exposures, the bay 

scallop was killed at 5 ug/L. 

NATIONAL CRITERIA: 

The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Uses indicate that, except possibly where a 

locally important species is very sensitive, freshwater aquatic 

organisms and their uses should not be affected unacceptably if 

the 4-day average concentration (in ug/L) of copper does not 

exceed the numerical value given, by .(0.8545[ ln(hardness) 3 -1.465) 

more than once every 3 years on the average and if the 1-hour 

average concentration (in ug/L) does not exceed the numerical 

value given by ,(0.9422 [ ln(hardness) 1-1.464) more than once every 

3 years on the average. For example, at hardnesses of 50, 100, 

and 200 mg/L as CaC03 the 4-day average concentrations of copper 

are 6.5, 12, and 21 ug/L, respectively, and the 1-hour average 

concentrations are 9.2, 18, and 34 ug/L. 

The procedures described in the Guidelines indicate that, 

except possibly where a locally important species is very 

sensitive, saltwater aquatic organisms and their uses should not 

be affected unacceptably if the 1-hour average concentration of 

copper does not exceed 2.9 ug/L more than once every 3 years on 

the average. 

EPA believes that a measurement such as tsacid-solublets would 
provide a more scientifically correct basis upon which to 

establish criteria for metals. The criteria were developed on 

this basis. However, at this time, no EPA approved methods for 

such a measurement are available to implement the criteria 

through the regulatory programs of the Agency and the States. 

The Agency is considering development and approval of methods 

for a measurement such as acid -soluble. Until available, 

however, EPA recommends applying the criteria using the total 

recoverable method. This has two impacts: (1) certain species of 

some metals cannot be analyzed directly because the total 

recoverable method does not distinguish between individual 

oxidation states, and (2) these criteria may be overly protective 

when based on the total recoverable method. 

The recommended exceedence frequency of 3 years is the 

Agency's best scientific judgment of the average amount of time 

it will take an unstressed system to recover from a pollution 

event in which exposure to copper exceeds the criterion. A 

stressed system, for example, one in which several outfalls occur 

in a limited area, would be expected to require more time for 

recovery. The resilience of ecosystems and their ability to 

recover differ greatly, however, and site -specific criteria may 

be established if adequate justification is provided. 

The use of criteria in developing waste treatment facilities 

requires the selection of an appropriate wasteload allocation 

model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

impractical, in which case one should rely on a steady -state 

model. The Agency recommends the interim use of 1Q5 or lQlO for 

0 
Criterion Maximum Concentration design flow and 745 or 7410 for 

the Criterion Continuous Concentration (CCC) design flow in 

steady-state models for unstressed and stressed systems 

respectively. These matters are discussed in more detail in the 

Technical Support Document for Water Quality-Based Toxics Control 

(U.S. EPA, 1985). 

HUMAN HEALTH CRITERIA: 

Sufficient data is not available for copper to derive a level 

which would protect against the potential toxicity of this 

compound. Using available organoleptic data, for controlling 

undesirable taste and odor quality of ambient water, the 

estimated level is 1 mg/L. It should be recognized that 

organoleptic data as a basis for establishing a water quality 

criteria have 1 imitations and have no demonstrated relationship 

to potential adverse human health effects. 

I 

(45 F.R. 79318 Nov. 28,1980) (50 F.R. 30784, July 29, 1985) 

SEE APPENDIX A FOR METHODOLOGY 
CYANIDE 

AQUATIC - LIFE SUMMARY: 

Data on the acute toxicity of free cyanide (the sum of 

cyanide present as HCN and CN-, expressed as CN) are available 

for a wide variety of freshwater species that are involved in 

diverse community functions. The acute sensitivities ranged from 

44.73 ug/L to 2,490 ug/L, but all of the species with acute 

sensitivities above 400 ug/L were invertebrates. A long-term 

survival, and a partial and life - cycle test with fish gave 

chronic values of 13.57, 7.849, and 16.39 ug/L, respectively. 

Chronic values for two freshwater invertebrate species were 

18.33 and 34.06 ug/L. Freshwater plants were affected at cyanide 

concentrations ranging from 30 ug/L to 26,000 ug/L. 

The acute toxicity of free cyanide to saltwater species ranged 

from 4.893 ug/L to >10,000 ug/L and invertebrates were both the 

most and least sensitive species. Long-term survival in an early 

life -stage test with the sheepshead minnow gave a chronic value 

of 36.12 ug/L. Long-term survival in a mysid life - cycle test 

resulted in a chronic value of 69.71 ug/L. Tests with the red 

macroalga, Champia parvula --I showed cyanide toxicity at 11 to 25 

ug/L, but other species were affected at concentrations up to 

3,000 ug/L. 

NATIONAL CRITERIA: 

The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses indicate that, except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 
unacceptably if the 4-day average concentration of cyanide does 

not exceed 5.2 ug/L more than once every 3 years on the average 

and if the 1-hour average concentration does not exceed 22 ug/L 

more than once every 3 years on the average. 

The procedures described in the Guidelines indicate that, 

except possibly where a locally important species is very 

sensitive, saltwater aquatic organisms and their uses should not 

be affected unacceptably if the 1-hour average concentration of 

cyanide does not exceed 1.0 ug/L more than once every 3 years on 

the average. 

EPA believes that a measurement such as "acid soluble" would 

provide a more scientifically correct basis upon which to 

establish criteria for cyanide. The criteria were developed on 

this basis. However, at this time, no EPA-approved methods for 

such a measurement are available to implement the criteria 

through the regulatory programs of the Agency and the States. 

The Agency is considering development and approval of methods 

for a measurement such as acid soluble. Until available, 

however, EPA recommends applying the criteria using the total 

recoverable method. These criteria may be overly protective when 

based on the total recoverable method. 

The recommended exceedence frequency of 3 years is the 

Agency's best scientific judgment of the average amount of time 

it will take an unstressed system to recover from a pollution 

event in which exposure to cyanide exceeds the criterion. A 

stressed system, for example, one in which several outfalls occur 

in a limited area, would be expected to require more time for 

recovery. The resilience of ecosystems and their ability to 
recover differ greatly, however, and site-specif ic criteria may 

be established if adequate justification is provided. 

The use of criteria in designing waste treatment facilities 

requires the selection of an appropriate wasteload a1 location 

model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

impractical, in which case one should rely on a steady-state 

model. The Agency recommends the interim use of 1Q5 or lQlO for 

Criterion Maximum Concentration design flow and 745 or 7410 for 

the Criterion Continuous Concentration design flow in steadystate models for unstressed and stressed systems respectively. 

These matters are discussed in more detail in the Technical 

Support Document for Water Quality-Based Toxics Control (U.S. 

EPA, 1985). 

H" HEALTH CRITERIA 

The ambient water quality criterion for cyanide is 

recommended to be identical to the existing drinking water 

standard which is 200 ug/L. Analysis of the toxic effects data 

resulted in a calculated level which is protective of human 

health against the ingestion of contaminated water and 

contaminated aquatic organisms. The calculated value is 

comparable to the present standard. For this reason a selective 

criterion based on exposure solely from consumption of 6.5 grams 

of aquatic organisms was not derived. 

NOTE: The U.S. EPA is currently developing Acceptable Daily 

Intake (ADI) or Verified Reference Dose (RfD) values for 

Agency-wide use for this chemical. The new value should 

be substituted when it becomes available. The January, 

1986, draft Verified Reference Dose document cites an RfD 

of .02 mg/kg/day for free cyanide. 

0 

-, 
CRITERIA: 

_. 

DDT 

-DDT AND - METABOLITES 

Aquatic Life 

For DDT and its metabolites the criterion to protect 

freshwater aquatic life as derived using the Guidelines is 0.0010 

ug/L as a 24-hour average and the concentration should not exceed 

1.1 ug/L at any time. 

For DDT and its metabolites the criterion to protect' 

saltwater aquatic life as derived using the Guidelines is 0.0010 

ug/L as a 24-hour average and the concentration should not exceed 

0.13 ug/L at any time. 

TDE - 

The available data for TDE indicate that acute toxicity to 

freswater aquatic life occurs at concentrations as low as 0.6 

ug/L and would occur at lower concentrations among species that 

are more sensitive than those tested. No data are available 

concerning the chronic toxicity of TDE to sensitive freshwater 

aquatic life. 

The available data for TDE indicate that acute toxicity to 

saltwater aquatic life occurs at concentrations as low as 3.6 

ug/L and would occur at lower concentrations among species that 

are more sensitive than those tested. No data are available 

concerning the chronic toxicity of TDE to sensitive saltwater 

aquatic life. 

- DDE 

The available data for DDE indicate that acute toxicity 

'./ 

to freshwater aquatic life occurs at concentrations as low as 
1,050 ug/L and would occur at lower concentrations among species 

that are more sensitive than those tested. No data are available 

concerning the chronic toxicity of DDE to sensitive freshwater 

aquatic life. 

The available data for DDE indicate that acute toxicity 

to saltwater aquatic life occurs in concentrations as low as 14 

ug/L and would occur at lower concentrations among species that 

are more sensitive than those tested. No data are available 

concerning the chronic toxicity of DDE to sensitive saltwater 

aquatic life. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to DDT through ingestion of 

contaminated water and contaminated aquatic organisms, the 

ambient water concentration should be zero, based on the 

nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at and The 

corresponding recommended criteria are 0.24 nq/L, 0.024 ng/L, and 

0.0024 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption 

of water, the levels are 0.24 ng/L, 0.024 ng/L, and 0.0024 

ng/L, respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
a CRITERION: 

DEMETON 

0.1 ug/L for freshwater and marine aquatic life 

RATIONALE: 

Static LC50 bioassays yielded toxicity values for the organophosphorus pesticide demeton for carp, goldfish, fathead minnow, 

channel catfish, guppy, rainbow, trout and bluegill, ranging from 

70 ug/L to 15,000 ug/L (Henderson and Pickering, 1958; Ludemann 

and Neumann, 1982; Macek and McAllister, 1970; McCann and Jasper, 

1972; Pickering et al. 1962). Results of these tests demonstrate 

an apparent sharp division in species sensitivity, with bluegill 

(Lepomis macrochirusr, rainbow trout (Salmo gairdneri) and 

guppy, (Poecilia --- --- reticulatg,' ------ being susceptible to lower 

concentrations while the remaining species were comparatively 

resistant. In the 96-hOUr exposures toxicity did not increase 

significantly with time, indicating that concentrations close to 

nominal may not have been maintained for more than a few hours, 

Bluegills with a 24-hour LC50 of 70 ug/L were the most sensitive 

fish (Mccann and Jasper, 1972). 

When fish were exposed to acutely toxic levels of demeton €or 

12 hours by Weiss (1959, 1961) the maximum inhibition of brain 

acetiylcholinesterase (AChE) was not reached. The lowest levels 

of AChE occurred after 24 to 48 hOUrS. It was demonstrated that 

maximum inhibition could last as long as two weeks after 

exposure, and subsequent recovery to levels approaching normal 

took many more weeks. Weiss (1958) reported a significant 

increase in mortality of fathead minnows exposed for a second 
time to the organophosphate, Sarin, before the fish had recovered 

normal brain AChE levels. The resistance of fully recovered 

fish was equal to that of previously unexposed controls. Weiss 

and Gakstatter (1964a) reported no significant inhibition of 

brain AChE in bluegills, goldfish and shiners (Notemigonus 

__ crysoleucasl, __ - ----- following 15-day exposures to demeton at 

continuously replenished, nominal concentrations of 1 ug/L. 

Acute toxicity values reported for invertebrates range from 

10 to 100,000 ug/L (Ludemann and Neumann, 1962; Sanders, 1972). 

In general, molluscs and tubifex worms were very resistant while 

the smaller crustaceans and insect larvae were susceptible. 

Ludemann and Neumann (1962) reported that Chironomus plumosus 

larvae were the most sensitive species they tested. A 24-hour 

exposure at 10 ug/L produced undefined effects while 100 percent 

were killed at 1000 ug/L. Calculated LC50 data for invertebrates 

apparently are limited to a single, nominal concentration static 

exposure of Gammarus ______ fasciatus (Sanders, 1972). These 24- and 

96-hourLC50 valuesarereportedas 500 and27 ug/L, indicatinga 

time-related effect not observed in the bioassays with fishes. 

As only a fewofthe sensitive specieshavebeentestedandgreat 

variance in response can result with different test methods, 

caution must be exercised in estimating the sub-acute 

concentration for aquatic fauna in general. It appears that no 

study has been made of possible residual effects other than AChE 

inhibition, which might result from short exposures to subacute 

concentrations of organophosphates. 

There are few data on the toxicity of demeton to marine 

organisms. Butler (1964) reported a 48-hour EC50 of 63 ug/L for 
the pink shrimp, Peneaus duorarum, and a 24-hour LC50 of 550 ug/L 

for the spot, Leiostomus xanthurus. 

Chronic demeton toxicity data for freshwater organism are not 

currently available. Since no data are available at this time to 

indicate long-term no-effect levels for aquatic organisms, a 

criterion must be derived based partly on the fact that all 

organophosphates inhibit the production of the AChE enzyme. 

Demeton is unique, however, in that the persistence of its AChEinhibiting ability is greater than that of 10 other common 

organophosphates, even though its acute toxicity is apparently 

less. The effective "half-life" of AChE inhibition for demeton 

is greater than one year (Weiss and Gakstatter, 1964b). Because 

such inhibition may be additive with repeated exposures and may 

be compounded by any of the organophosphates, it is recommended 

that a criterion for demeton be based primarily on its enzymeinhibiting potential. A criterion of 0.1 ug/L demeton for 

freshwater and marine aquatic life is recommended since it will 

not be expected to significantly inhibit AChE over a prolonged 

period of time. In addition, the criteria recommendation is in 

close agreement with the criteria for the other organophosphates. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
DICHLOROBENZENES 

Aquatic - Life 

The available data for dichlorobenzenes indicate that acute 

and chronic toxicity to freshwater aquatic life occur at 

concentrations as low as 1,120 and 763 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for dichlorobenzenes indicate that acute 

toxicity to saltwater aquatic life occurs at concentrations as 

low as 1,970 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic' toxicity of dichlorobenzenes to 

0 sensitive saltwater aquatic life. 

Human Health 

For the protection of human health from the toxic properties 

of dichlorobenzene ingested through water and contaminated 

aquatic organisms, the ambient water criterion is determined to 

be 400 ug/L. 

For the protection of human health from the toxic properties 

of dichlorobenzenes ingested through contaminated aquatic 

organisms alone, the ambient water criterion is determined to be 

2.6 mg/L. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B.FOR METHODOLOGY 0 ., 
DICHLOROBENZIDINE 

CRITERIA: 

Aquatic Life 

The data base available for dichlorobenzidines and freshwater 

organisms is limited to one test on bioconcentration of 3,3- 

dichlorobenzidine, and therefore, no statement can be made 

concerning acute or chronic toxicity. 

No saltwater organisms have been tested with any 

dichlorobenzidine, and therefore, no statement can be made 

concerning acute or chronic toxicity. 

Human Health 

For the maximum protection of human health from the 

potential carcinogenic effects of exposure to dichlorobenzidine 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at and lo-'. 

The corresponding recommended criteria are 0.103 ug/L, 0.010 

ug/L, and 0.001 ug/L, respectively. If these estimates are 

made for consumption'of aquatic organisms only, excluding 

consumption of water, the levels are 0.204 ug/L, 0.020 ug/L, and 

0.002 ug/L, respectively. 

0 

(45 F.R. 79318, November 28, 1980) 

_.* SEE APPENDIX B FOR METHODOLOGY 
DICHLOROETHYLENES 

Aquatic Life 

The available data for dichloroethylenes indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 11,600 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No definitive 

data are available concerning the chronic toxicity of 

dichloroethylenes to sensitive freshwater aquatic life. 

The available data for dichloroethylenes indicate that acute 

and chronic toxicity to saltwater aquatic life occurs at 

concentrations as low as 224,000 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity of 

dichloroethylenes to sensitive saltwater aquatic life. 

Human Health 

1,1-Dichloroethylene 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to 1,l dichloroethylene through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, 

based on the non threshold assumption for this chemical. 

However, zero level may not be attainable at the present time. 

Therefore, the levels which may result in incremental increase of 

cancer risk over the lifetime are estimated at 10-5e 10- 0 .- 6, and The corresponding recommended criteria are 

0.33 ug/L, 0.033 ug/L, and 0.003 ug/L, respectively. If 
these estimates are made for consumption of aquatic 

organisms only, excluding consumption of water, the levels are 

18.5 ug/L, 1.85 ug/L, and 0.185 ug/L, respectively. 

1,2-Dichloroethylene 

Using the present guidelines, a satisfactory criterion cannot 

be derived at this tine because of insufficient available data 

for 1,2-dichloroethylene. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
2,4-DICHLOROPHENOL 

CRITERIA: 

Aquatic Life 

The available data for 2,4-dichlorophenol indicate that acute 

and chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 2,020 and 365 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. Mortality to early life stages of 

one species of fish occurs at concentrations as low as 70 ug/L. 

Only one test has been conducted with saltwater organisms and 

2,4-dichlorophenol and therefore, no statement can be made 

concerning acute or chronic toxicity. 

Human Health 

For comparison purposes, two approaches were used to derive 

criterion levels for 2,4-dichlorophenol. Based on available 

toxicity data, to protect public health the derived level is 3.09 

mg/L. Using available organoleptic data, to control undesirable 

taste and odor qualities of ambient water the estimated level is 

0.3 ug/L. It should be recognized that organoleptic data have 

limitations as a basis for establishing a water quality 

criterion, and have no demonstrated relationship to potential 

adverse human health effects. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
DICHLOROPROPANES/DICHLOROPROPROPENES 

CRITERIA: 

Aquatic Life 

The available data for dichloropropanes indicate that acute 

and chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 23,000 and 5,700 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for dichloropropene indicate that acute 

and chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 6,060 and 244 ug/L, respeptively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for dichloropropane indicate that acute 

and chronic toxicity to saltwater aquatic life occur at 

concentrations as low as 10,300 and 3,040 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for dichloropropene indicate that acute 

toxicity to saltwater aquatic life occurs at concentrations as 

low as 790 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data 

are available concerning the chronic toxicity of 

dichloropropene to sensitive saltwater aquatic life. 
Human Health 

Using the present guidelines, a satisfactory criterion cannot 

be derived at this time because of insufficient available data 

for dichloropropanes. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
Aquatic Life 

The available data for 2,4-dimethylphenol indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 2,120 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of dimethylphenol 

to sensitive freshwater aquatic life. 

No saltwater organisms have been tested with 2,4- 

dimethyl-phenol and therefore, no statement can be made 

concerning acute or chronic toxicity. 

Human Health 

Sufficient data are not available for 2,4-dimethylphenol to 

derive a level which would protect against the potential toxicity 

of this compound. Using available organoleptic data, to control 

undesirable taste and odor quality of ambient water the estimated 

level is 400 ug/L. It should be recognized that organoleptic 

data have limitations as a basis for establishing a water quality 

criterion, and have no demonstrated relationship to 

potential adverse human health effects. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERIA: 

DINITROTOLUENE 

Aquatic Life 

The available data for dinitrotoluenes indicate that acute 

and chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 330 and 230 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for dinitrotoluenes indicate that acute 

toxicity to saltwater aquatic life occurs at concentrations as 

low as 590 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of dinitrotoluenes 

to sensitive saltwater aquatic life but a decrease in algal cell 

numbers occurs at concentrations as low as 370 ug/L. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to 2,4-dinitrotoluene through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentration should be zero, based 

on the nonthreshold assumption for this chemical. However, 

zero level may not be attainable at the present time. Therefore, 

the levels which may result in incremental increase of cancer 

risk over the lifetime are estimated at and The 

corresponding recommended criteria are 1.1 ug/L, 0.11 ug/L, and 

0.011 ug/L, respectively. If these estimates are made for 

consumption of aquatic organ-isms only, excluding consumption 
of water, the levels are 91 ug/L, 9.1 ug/L, and 0.91 ug/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
DIPHENYLHYDRAZINE 

Aquatic Life 

The available data for 1,2-diphenylhydrazine indicate that 

acute toxicity to freshwater aquatic life occurs at 

concentrations as low as 270 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity of 

1,2-diphenylhydrazine to sensitive freshwater aquatic life. 

NO saltwater organisms have been tested with 1,2- 

diphenylhydrazine and therefore, no statement .can be made 

concerning acute or chronic toxicity. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to diphenylhydrazine through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations shopld be zero, based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at and 

The corresponding recommended criteria are 422 ng/L, 42 ng/L, and 

4 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 5.6 ug/L, 0.56 ug/L, and 0.056 ug/L, 

0 

respectively. 0 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
ENDOSULFAN 

Aquatic Life 

For endosulfan the criterion to protect freshwater aquatic 

life as derived using the Guidelines is 0.056 ug/L as a 24-hour 

average and the concentration should not exceed 0.22 ug/L at any 

time. 

For endosulfan the criterion to protect saltwater 

aquatic life as derived using the Guidelines is 0.0087 ug/L as a 

24-hour average and the concentration should not exceed 0.034 

ug/L at any time. 

Human Health 

For the protection of human 'health from the toxic properties 

of endosulfan ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 74 

ug/L. 

For the protection of human health from the toxic properties 

of endosulfan ingested through contaminated aquatic 

organisms alone, the ambient water criterion is determined to be 

159 ug/L. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERIA: 

*ENDRIN 

Aquatic Life 

For endrin the criterion to protect freshwater aquatic life 

as derived using the Guidelines is 0.0023 ug/L as a 24-hour 

average, and the concentration should not exceed 0.18 ug/L at any 

time. 

For endrin the criterion to protect saltwater aquatic life as 

derived using the Guidelines is 0.0023 ug/L as a 24-hour average, 

and the concentration should not exceed 0.037 ug/L at any time. 

Human Health 

The ambient water quality criterion for endrin is recommended 

to be identical to the existing water standard which is 1.0 ug/L. 

Analysis of the toxic effects data resulted in a calculated level 

which is protective of human health against the ingestion of 

contaminated water and contaminated aquatic organisms. The 

calculated value is comparable to the present standard. For 

this reason a selective criterion based on exposure solely from 

assumption of 6.5 g of aquatic organisms was not derived. 

*Indicates suspended, canceled or restricted by W.S. EPA Office 

of Pesticides and Toxic Substances 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
0 CRITERIA: 

The avai 3 B da 

ETHYLBENZENE 

Aquatic Life 

3 for ethylbenzene idicate t at acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 32,000 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No definitive 

data are available concerning the chronic toxicity of 

ethylbenzene to sensitive freshwater aquatic life. 

The available data for ethylbenzene indicate that 

acute toxicity to saltwater aquatic life occurs at 

concentrations as low as 430 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity 

of ethylbenzene to sensitive saltwater aquatic life. 

Human Health 

For the protection of human health from the toxic properties 

of ethylbenzene ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 1.4 

W/L. 

For the protection of human health from the toxic properties 

of ethylbenzene ingested through contaminated aquatic organisms 

alone, the ambient water criterion is determined to be 3.28 mg/L. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 

NOTE: The U.S. EPA is currently developing Acceptable Daily 

Intake (ADI) or Verified Reference Dose (RfD) values for 

Agency-wide use for this chemical. The new value should 

be substituted when it becomes available. The January, 

1986, draft Verified Reference Dose document cites an RfD 

of 0.1 mg/kg/day for ethylbenzene. 
FLUORANTHENE 

Aquatic Life 

The available data for fluoranthene indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 3,980 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of f luoranthene to 

sensitive -freshwater aquatic life. 

The available data €or fluoranthene indicate that acute and 

chronic toxicity to saltwater aquatic life occur at 

concentrations as low as 40 and 16 ug/L, respectively, and would 

occur at lower concentrations' among species that are more 

a sensitive than those tested. 

Human Health 

For the protection of human health from the toxic properties 

of fluoranthene ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 42 

ug/L. 

For the protection of human health from the toxic properties 

of fluoranthene ingested through contaminated aquatic organisms 

alone, the ambient water criterion is determined to be 54 ug/L. 

(45 F.R. 79318, Novembe-r 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
GASES, TOTAL DISSOLVED 

CRITERION: 

To protect freshwater and marine aquatic life, the total 

dissolved gas concentrations in water should not exceed 110 

percent of the saturation value for gases at the existing 

atmospheric and hydrostatic pressures. 

RATIONALE: 

Fish in water containing excessive dissolved gas pressure or 

tension are killed when dissolved gases in their circulatory 

system come out of solution to form bubbles (emboli) which block 

the flow of blood through the capillary vessels. In aquatic 

organisms this is commonly referred to as "gas bubble diseasett. 

External bubbles (emphysema) also appear in the fins, on the 

opercula, in the skin and in other body tissues. Aquatic 

invertebrates are also affected by gas bubble disease, but 

usua1l.y at supersaturation levels higher than those lethal to 

fish. 

0 

The standard method of analyzing for gases in solutions has 

been the Van Slyke method (Van Slyke et al. 1934); now, gas 

chromatography also is used for determination of individual and 

total gases. For determination of total gas pressure, Weiss has 

developed the saturometer, a device based upon a thin-wall 

silicone rubber tube that is permeable to gases but impermeable 

0 to water. Gases pass from the water through the tube, thus 

raising the internal gas pressure which is measured by a 
manometer or pressure gauge connected to the tube (NAS, 1974). 

This method alone does not separate the total gas pressure into 

the separate components, but Winkler oxygen determinations can be 

run simultaneously, and gas concentrations can be calculated. 

Total dissolved gas concentrations must be determined because 

analysis of individual gases may not determine with certainty 

that gas supersaturation exists. For example, water could be 

highly supersaturated with oxygen, but if nitrogen were at less 

than saturation, the saturation as measured by total gas pressure 

might not exceed 100 percent. Also, if the water was highly 

supersaturated with dissolved oxygen, the oxygen alone might be 

sufficient to create gas pressures or tensions greater than the 

Criterion limits, but one would not know the total gas pressure 

or tension, or by how much the criterion was exceeded. The rare 

and inert gases such as argon, neon and helium are not usually 

involved in causing gas bubble disease as their contribution to 

total gas pressures is very low. Dissolved nitrogen (NZ), which 

comprises roughly 80 percent of the earth's atmosphere, is nearly 

inert biologically and is the most significant cause of gas 

bubble disease in aquatic animals. Dissolved oxygen, which is 

extremely bioactive, is consumed by the metabolic processes of 

the organism and is less important in causing serious gas bubble 

disease though it may be involved in initiating emboli formation 

in the blood (Nebeker et al. 1976a). 

Percent saturation of water containing a given amount of gas 

varies with the absolute temperature and with the pressure. 

Because of the pressure changes, percent saturation with a given 
amount of gas changes with depth of the water. Gas 

supersaturation decreases by 10 percent per meter of increase in 

water depth because of hydrostatic pressure; a gas that is at 130 

percent saturation at the surface would be at 100 percent 

saturation at 3 meters' depth. Compensation €or altitude may be 

needed because a reduction in atmospheric pressure changes the 

water/gas equilibria, resulting in changes in solubility of 

dissolved gases. 

There are several ways that total dissolved gas 

supersaturation can occur: 

1. Excessive biological activity--dissolved oxygen 

concentrations often reach supersaturation because of excessive 

algal photosynthesis. Renfro (1963) reported gas bubble disease 

in fishes resulting, in part, from algal blooms. Algal blooms 

often accompany an increase in water temperature and this higher 

temperature further contributes to supersaturation. 

I 

2. Lindroff (1957) reported that water spillage at 

hydropower dams caused supersaturation. When excess water is 

spilled over the face of a dam it entrains air as it plunges to 

the stilling or plunge pool atthebaseofthedam. Themomentum 

of the fall carries the water and entrained gases to great depths 

in the pool; and, under increased hydrostatic pressure, the 

entrained gases are driven into solution, causing supersaturation 

of dissolved gases. 

3. Gas bubble disease may be induced by discharges from 

power-generating and other thermal sources (Marcello et al. 

./ 

1975). Cool, gas-saturated water is heated as it passes through 

the condenser or heat exchanger. As the temperature of the water 
rises, percent saturation increases because of the reduced 

solubility of gases at higher temperatures. Thus, the discharged 

water becomes supersaturated with gases and fish or other 

organisms living in the heated water may exhibit gas bubble 

disease (DeMont and Miller, 1972: Malouf et al. 1972; Keup, 

1975). ! 

In recent years, gas bubble disease has been identified as a 

major problem affecting valuable stocks of salmon and trout in 

the Columbia River system (Rulifson and Abel, 1971). The disease 

is caused by high concentrations of dissolved atmospheric gas 

which enter the river's water during heavy spillinq at 

hydroelectric dams. A report by Ebel et al. (1975) presents 

results from field and laboratory studies on the lethal, 

sublethal and physiological effects of gas on fish, depth 

distribution of fish in the river (fish can compensate for some 

high concentrations of gas by moving deeper into the water 

column), detection and avoidance of gas concentrations by fish, 

intermittent exposure of fish to gas concentrations, and 

bioassays of many species of fish exposed to different 

concentrations of gas. Several conclusions resulting from these 

studies are: 

1. When either juvenile or adult salmonids are confined to 

shallow water (1 m), substantial mortality occurs at and above 

115 percent total dissolved gas saturation. 

2. When either juvenile or adult salmonids are free to sound 

and obtain hydrostatic compensation either in the laboratory or 

in the field, substantial mortality still occurs when saturation , 
levels (of total dissolved gases) exceed 120 percent saturation. 

3. On the basis of survival estimates made in the Snake 

River from 1966 to 1975, it is concluded that juvenile fish 

losses ranging from 40to 95 percentdooccurandamajorportion 

of this mortality can be attributed to fish exposure to 

supersaturation by atmospheric gases during years of high flow. 

4. Juvenile salmonids subjected to sublethal periods of 

exposure to supersaturation can recover when returned to normally 

saturated water, but adults do not recover and generally die from 

direct and indirect effects of the exposure. 

5. Some species of salmon and trout can detect and avoid 

supersaturated water; others may not. 

6. Higher survival was 'observed during periods of 

intermittent exposure than during continuous exposure. 

7. In general, in acute bioassays, salmon and trout were 

less tolerant than the nonsalmonids. 

Dawley and Ebel (1975) found that exposure of juvenile spring 

chinook salmon, Oncorhynchus tshawytscha, -- and steelhead trout, 

Salmo qairdneri, to 120 percent saturation for 1.5 days resulted 

in over 50 percent mortality; 100 percent mortality occurred in 

less than 3 days. They also determined that the threshold level 

where significant mortalities begin occurring is at 115 percent 

nitrogen saturation (111 percent total gas saturation in this 

test). 

Rucker (1974), using juvenile coho salmon, Oncorhynchus 

kisutch, determined the effect of individual ratios of oxygen and 

nitrogen and established that a decrease in lethal effect -., 

occurred when the nitrogen content fell below 109 percent 
saturation even though total gas saturation remained at 119 

percent saturation, indicating the importance of determining the 

concentration of the individual components (02 and N2) of the 

atmospheric supersaturation. Nebeker et al. (1976a), using 

juvenile sockeye salmon, Oncorhynchus nerka, also showed that 

there was a significant increase in fish mortality when the 

nitrogen concentration was increased while holding the total 

percent saturation constant. They also showed that there was no 

significant difference in fish mortality at different C02 

concentrations. 

Research collected by Bouck et al. (1975) showed that gas 

supersaturated water at and above 115 percent total gas 

saturation is acutely lethal to most species of salmonids, with 

120 percent saturation and above rapidly lethal to all salmonids 

tested. Levels as low as 110 percent will produce emphysema in 

most species. Steelhead trout were most sensitive to gassupersaturatea water followed by sockeye salmon, Oncorhyncnus 

nerka. Chinook salmon, ______ Oncorhynchus _____ -_--_ tshawytscha -----I were 

intermediate in sensitivity. Coho salmon, Oncorhyncnus kisutch, 

were significantly the more tolerant of the salmonids though 

----- 

still much more susceptible than non-salmonids like bass or carp. 

- DaFnia -- rmqna __ exhibited a sensitivity to supersaturation 

similar to that of the salmonids (Nebeker et al. 1975), with 115 

percent saturation lethal within a few days. Stoneflies exhibited 

an intermediate sensitivity similar to bass with mortality at 130 

percent saturation. Crayfish were very tolerant, with levels 

near 140 percent total gas saturation resulting in mortality. 
No differences are proposed in the criteria €or freshwater 

and marine aquatic life as the data available indicate that there 

probably is little difference in overall tolerances between 

marine and freshwater species. 

The development of gas bubble disease in menhaden, Soortia 

sp., and their tolerance to gas saturation in laboratory 

bioassays and in the field (Pilgrim Nuclear Power Station 

Discharge Canal) are discussed by Clay et al. (1975) and Marcello 

et al. (1975). At 100 percent and 105 percent nitrogen 

saturation, no gas bubbles developed externally or in any of the 

internal organs of menhaden. At 105 percent nitrogen saturation, 

however, certain behavioral changes became apparent. Fish 

sloughed of€ mucus, swam erratically, were more excitable, and 

became darker in color. Menhaden behavioral changes observed at 

110 percent nitrogen saturation were similar to those noted at 

3.05 percent. In addition, at 110 percent gas emboli were found 

in the intestines, the pyloric caeca, and occasionally the 

operculum. The behavioral changes described were also observed 

at 115 percent, and clearly defined subcutaneous emphysema was 

observed inthe fins and occasionally inthe eye. At120percent 

and 130 percent nitrogen saturation, menhaden developed within a 

few hours classic symptoms of gas bubble disease. Externally, 

emboli were evident in all fins, the operculum and within the 

oral cavity. 

Exophthalmia also occurred and emboli developed in internal 

The bulbous arteriosis and swim bladder were severely 

J distended, and emboli were found along the length of the gill 

arterioles, resulting in hemostasis. At water temperatures of 30 

organs. 0 
OC, menhaden did not survive, regardless of gas saturation level. 

At water temperatures of 15 , 22 , and 25 OC 100 percent of the 

menhaden died within 24 hours at 120 percent and 130 percent gas 

saturation. Fifty percent died after 96 hours at 115 percent (22 

OC) Menhaden survival after 96 hours at 110 percent nitrogen 

saturation ranged from 92 percent at 22O and 25' to 83 percent 

at 15 OC. Observations on the relationship between the mortality 

rate of menhaden and gas saturation levels at Pilgrim Station 

during the April 1975, incident suggest that the fish may 

tolerate somewhat higher gas saturation levels in nature. 

It has been shown by Bouck et al. (1975) and Dawley et al. 

(1975) that survival of salmon and steelhead smolts in seawater 

is not affected by prior exposure to gas supersaturation while in 

fresh water. No significant mortality of juvenile coho and 

sockeye salmon occurred when they were exposed to sublethal 

concentrations of supersaturated water and then transferred to 

seawater (Nebeker et al. 197633). 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLLIGY 
GUTHION 

CRITERION: 

.01 ug/L for freshwater and marine aquatic life. 

RATIONALE : 

Ninety-six-hour LC50 values for fish exposed to the 

organophosphorus pesticide guthion range from 4 to 4270 ug/L 

(Katz, 1961: Pickering et al. 1962; Lahav and Sarig, 1969; Macek 

et al. 1969; Macek and McAllister, 1970). The only long-term 

fish exposure data available are those obtained recently by 

Adelman and Smith (unpublished data) Decreased spawning (eggs 

produced per female) was observed in fathead minnows, PimephalE 

prrelas ---I exposed during a complete life cycle. An estimated 

"safe" long-term exposure concentration for fathead minnows lies 

between 0.3 and 0.5 ug/L. survival of larvae was reduced at 

approximately 0.7 ug/L. 

0 

An investigation of the persistence of guthion in fish 

revealed that 50 percent of the chemical was lost in less than 

one week (Meyer, 1965). Analysis of plankton and pond water in 

the same study indicated a 50 percent loss of guthioninabout48 

hours. Flint et al. (1970) determined the half-life of guthion 

at 30C in pond water and in a phosphate buffer protected from 

light in the laboratory. The half-life in pond water was 1.2 

days whereas that in the laboratory solution was 10 days. The 

more rapid degradation in pond water was attributed to the effect 

of sunlight and microorganisms. 

Organophosphate pesticides are toxic because they inhibit the 
enzyme acetylcholinesterase (AChE) which is essential to nerve 

impulse conduction and transmission (Holland et al. 1967). Weiss 

(1958, 1959, 1961) demonstrated that a 40 to 70 percent 

inhibition of fish brain AChE usually is lethal. Centrarchids 

generally are considered one of the more sensitive groups of fish 

to guthion (Pickering et al. 1962; Weiss and Gakstatter, 1964; 

Meyer, 1965). Weiss and Gakskatter (1964) found that over a 15- 

day period bluegills, -_ Lepomis __ macrochirus 1 exhibited AChE 

inhibition at 1.0 ug/L guthion but not at 0.1 ug/L. Exposure at 

0.05 ug/L for 30 days also failed to produce inhibition below the 

range of normal variation, but the authors stated that it 

appeared there was a downward trend in brain enzyme activity and 

that if exposure was continued a definite reduction might 

develop. Weiss (1961) found that about 30 days were required for 

fathead minnow and bluegill brain AChE levels to recover after 8 

to 24 hours exposure to 10 ug/L guthion. 

Benke and Murphy (1974) showed that repetitive injection of 

fish with guthion caused cumulative inhibition of brain AChE and 

mortality. After substantial inhibition by guthion exposure, it 

takes several weeks for brain AChE of fishes to return to normal 

even though exposure is discontinued (Weiss, 1959, 1960; Carter, 

1971). Inhibition of brain AChE of fishes by 46 percent or more 

has been associated with harmful effects in exposures to there 

organophosphate pesticides for a life cycle (Eaton, 1970) and for 

shorter periods (Carter, 1971; Coppage and Duke, 1971; Coppage, 

1972; Coppage and Matthews, 1974: Post and Leasure, 1974; Coppage 

et al. in press). In static tests, similar inhibition of AChE 

and mortality were caused in the sheepshead minnow, Cyprinodon 
variegatus, in 2, 24, 48 and 72 hours at concentrations of 50, 7, 

3.5 and 3 ug/L, respectively (Coppage, 1972). These data 

indicate that reduction of brain AChE activity of marine fishes 

by 70 to 80 percent or more in short-term exposures to guthion 

may be associated with some deaths. 

There is no evidence to indicate that guthion would cause 

adverse effects through the food chain. Tissue residue 

accumulation for whole fish calculated from the data of Meyer 

(1965) indicate no more than a twentyfold accumulation. LC50 

toxicity values for birds are relatively high and range from 70 

to 2,000 mg/kg (Tucker and Crabtree, 1970). 

Ninety-six-hour LC50 values for aquatic invertebrates range 

from 0.10 to 22.0 ug/L (Nebeker and Gaufin, 1964; Gaufin et al. 

1965: Jensen and Gaufin, 1966: Sanders and Cope, 1968: Sanders, 

1969, 1972). Sanders (1972) exposed the grass shrimp, 

Paleomonetes --_-__ kadiakensis r to guthion in a continuous flow 

bioassay for up to 20 days and found that the 5- and 20-day LC50 

values were 1.2 and 0.16 ug/L, respectively. He found that the 

amphipod, ---- Gammarus -I--_- fasciatus 1 was the most sensitive aquatic 

organism tested, with a 96-hour LC50 of 0.10 ug/L. Jensen and 

Gaufin (1966), also using a continuous flow system, exposed two 

species of stonefly naiads in 4- and 30-day studies. They 

observed 96-hour and 30-day LC50 values for Acroneuria pacifica 

of 2.0 and 0.24 ug/L, respectively, whereas for Pteronarcys - 

- californica the values were 4.6 and 1.3 ug/L, respectively. 

Results of other toxicity studies on marine organisms have 0 

1.3 been reported. The 24-hour LC50 for the white mullet, _- MUG1 _ 
curema, was found to be 5.5 ug/L guthion (Butler, 1963). The 96- 

hour LC50 for the striped mullet, Mugil cephalus, was determined 

by Lahav and Sarig (1969) to be 8 ug/L guthion. Portman (1972) 

reported the 48-hour LC50 for the fish, - Pleuronectes - limanda, to 

be 10 to 30 ug/L. The 48-hour LC50 for the European shrimp, 

Cranqon crangon, was found to be 0.33 ug/L guthion (Portman, 

1972). Butler (1963) found that the 24-hour EC50 €or blue crab, 

------- Callinectes - sapidus, was 550 ug/L and the 48-hour EC50 €or pink 

shrimp, Penaeus duorarum, was 4.4 ug/L guthion. The 48-hour TLm 

was estimated to be 620 ug/L €or fertilized oyster eggs, 

Crassostrea vixinica -- ---I and 860 ug/L for fertilized clam eggs, 

Mercenaria - mercenaria (Davis and Hidu, 1969). 

A criterion level of .01 ug/L for guthion is based upon use 

of an 0.1 application factor applied to the 96-hour LC50 of 0.1 

ug/L for -Gammarus ---- and a similar value of 0.3 ug/L for the 

European shrimp. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
HALOETHERS 

CRITERIA: 

Aquatic Life 

The available data for haloethers indicate that acute and 

chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 360 and 122 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

No saltwater organisms have been tested with any haloether 

and therefore, no statement can be made concerning acute or 

chronic toxicity. 

Human Health 

Using the present guidelines, a satisfactory criterion cannot 

be derived at this time because of insufficient available data 

for ha 1 oethers. 

0 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
HALOMETHANES 

CRITERIA: 

Aquatic Life 

The available data for halomethanes indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low a 11,000 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chroni-c toxicity of halomethanes to 

sensitive freshwater aquatic life. 

The available data for halomethanes indicate that acute and 

chronic toxicity to saltwater aquatic life occurs at 

concentrations as low as 12,000 and 6,400 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. A decrease in algal cell numbers 

occurs at concentrations as low as 11,500 ug/L. 

0 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to chloromethane, 

bromomethane, dichloromethane, bromodichloromethane, 

tribromomethane, dichlorodifluoromethane, trichlorofluoromethane, 

or combinations of these chemicals through ingestion of 

contaminated water and aquatic organisms, the ambient water 

concentration should be zero, based on the non threshold 

assumption for this chemical, However, zero level may not be 

attainable at the present time. Therefore, the levels which may 0 result in incremental increase of cancer risk over the lifetime - 
are estimated at loe6 and The corresponding 

recommended criteria are 1.9 ug/L, 0.19 ug/L, and 0.019 uq/L, 

respectively. If these estimates are made for consumption of 

aquatic organisms only, excluding consumption of water, the 

levels are 157 ug/L, 15.7 ug/L, and 1.57 ug/L, respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
HARDNESS 

I"RODUC"ION: 

Water hardness is caused by the polyvalent metallic ions 

disolved in water. In fresh water these are primarily calcium 

and magnesium although other metals such as iron, strontium and 

manganese contribute to the extent that appreciable 

concentrations are present. Hardness commonly is reported as an 

equivalent concentration of calcium carborate (CaC03). 

The concept of hardness comes from water supply practice. It 

is measured by soap requirements for adequate lather formation 

and as an indicator of the rate of scale formation in hot water 

heaters and low pressure boilers. A commonly used classification 

is given in the following table (Sawyer, 1960). 

TABLE 3. 

Classification of Water by Hardness Content 

Conc. mg/L CaC03 Description 

0 - 75 

75 - 150 

150 - 300 

300 and up 

soft 

moderately hard 

hard 

very hard 

Natural sources of hardness principally are limestones which 

are dissolved by percolating rainwater made acid by dissolved 

carbon dioxide. Industrial and industrially related sources 

include the inorganic chemical industry and discharges from 

operating and abandoned mines. 

Hardness in fresh water frequently is distinguished in 

carbonate and non-carbonate fractions. The carbonate fraction is 

chemically equivalent to the bicarbonates present in water. 

I 
Since bicarbonates generally are measured as alkalinity, the 

carbonate hardness usually is considered equal to the 

alkalinity. 

RATIONALE: 

The determination of hardness in raw waters subsequently 

treated and used for domestic water supplies is useful as a 

parameter to characterize the total dissolved solids present and 

for calculating dosages where lime-soda softening is practiced. 

Because hardness concentrations in water have not been proven 

health related, the final level achieved principally is a 

function of economics. Since hardness in water can be removed 

with treatment by such processes as lime-soda softening and 

zeolite or ion exchange systems, a criterion for raw waters used 

for public water supply is not practical. 

The effects of hardness on freshwater fish and other aquatic 

life appear to be related to the ions causing the hardness rather 

than hardness. Both the NTAC (1968) and NAS (1974) panels have 

recommended against the use of the term hardness but suggest the 

inclusion of the concentrations of the specific ions. This 

procedure should avoid confusion in future studies but is not 

helpful in evaluating previous studies. For most existing data, 

it is difficult to determine whether toxicity of various metal 

ions is reduced because of the formation of metallic hydroxides 

and carbonates caused by the associated increases in alkalinity, 

or because of an antagonistic effect of one of the principal 

cations contributing to hardness, e.g., calcium, or a combination 

of both effects. Stiff (1971) presented a theory (without proof) 
that if cupric ions were the toxic form of copper whereas copper 

carbonate complexes were relatively non-toxic, then the observed 

difference in toxicity of copper between hard and soft waters can 

be explained by the difference in alkalinity rather than 

hardness. Doudoroff and Katz (1953), in their review of the 

literature on toxicity, presented data showing that increasing 

calcium in particular reduced the toxicity of other heavy metals. 

Under usual conditions in fresh water and assuming that other 

bivalent metals behave similarly to copper, it is reasonable to 

assume that both effects occur simultaneously and explain the 

observed reduction of toxicity of metals in waters containing 

carbonate hardness. The amount of reduced toxicity related to 

hardness, as measured by a 40-hour LC50 for rainbow trout, has 

been estimated to be about four times for copper and zinc when 

the hardness was increased from 10 to 100 mg/L as CaC03 (NAS, 

1974) - 

Limits on hardness for industrial uses are quite variable. 

Table 4 lists maximum values that have been accepted by various 

industries as a source of raw water (NAS, 1974). Subsequent 

treatment generally can reduce hardness to tolerable limits 

although costs of such treatment are an important factor in 

determining its desirability €or a particular water source. 

Hardness is not a determination of concern for irrigation use 

of water. The concentrations of the cations calcium and 

magnesium, which comprise hardness, are important in determining 

the exchangeable sodium in a given water. This particular 

calculation will be discussed under total dissolved solids rather 
TABLE 4. 

Maximum Hardness Levels Accepted 

By Industry as a Raw Water Source* 

Industry 

Electric utilities 

Maximum Concentration 

m&L as CaC03 

- 

5,000 

Textile 120 

Pulp and paper 475 

Chemical 1,000 

Petroleum 900 

Primary metals 1,000 

* Requirements for final use within a process may be essential11 

zero, which requires treatment for concentration reductions. 
than hardness. a 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
HEFTACHLOR 

CRITERIA: 

Aquatic Life 

For heptachlor the criterion to protect freshwater aquatic 

life as derived using the Guidelines is 0.0038 ug/L as a 24-hour 

average, and the concentration should not exceed 0.52 ug/L at any 

time. 

For heptachlor the criterion to protect saltwater aquatic 

life as derived using the Guidelines is 0.0036 ug/L as a 24-hour 

average, and the concentration should not exceed 0.053 ug/L at 

any time. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to heptachlor through ingestion 

of contaminated water and contaminated aquatic organisms, the 

ambient water concentration should be zero, based on the non 

threshold assumption for this chemical. However, zero level 

may not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

lifetime are estimated at and The 

corresponding recommended criteria are 2.00 ng/L, 0.20 ng/L, and 

0.020 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 2.04 ng/L, 0.20 ng/L, and 0.020. ng/L, 

0 

respectively. 0 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
HEXACHLOROBUTADIENE 

CRITERIA: 

Aquatic Life 

The available data for hexachlorobutadiene indicate 

that acute and chronic toxicity to freshwater aquatic life 

occur at concentrations as low as 90 and 9.3 ug/L, respectively, 

and would occur at lower concentrations among species that are 

more sensitive than those tested. 

The available data for hexachlorobutadiene indicate that 

acute toxicity to saltwater aquatic life occurs at concentrations 

as low as ?2 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

I 

available concerning the chronic toxicity of hexachlorobutadiene 

0 to sensitive saltwater aquatic life. 

Human Health 

For the maximum protection of human health from the 

potential carcinogenic effects of exposure to 

hexachlorobutadiene through ingestion of contaminated water 

and contaminated aquatic organisms, the ambient water 

concentrations should be zero, based on the nonthreshold 

assumption for this chemical. However, zero level may not be 

attainable at the present time. Therefore, the levels which may 

result in incremental increase of cancer risk over the 

lifetime 'are estimated at and The 

corresponding recommended criteria are 4.47 ug/L, 0.45 0 i ug/L, and 0.045 ug/L, respectively. If these estimates are -. 

made for consumption of aquatic organisms only, excluding 
consumption of water, the levels are 500 ug/L, 50 ug/L, and 5.0 

ug/L, respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
HEXACHUIROCYCLOHEXANE 

CRITERIA: 

Aquatic Life 

Lindane 

For lindane the criterion to protect freshwater aquatic life 

as derived using the Guidelines is 0.080 ug/L as a 24-hour 

average and the concentration should not exceed 2.0 ug/L at any 

time. 

For saltwater aquatic life the concentration of lindane 

should not exceed 0.16 ug/L at any time. No data are available 

concerning the chronic toxicity of lindane to sensitive saltwater 

aquatic life. 

- BHC 

The available data for a mixture of isomers of BHC indicate 

that acute toxicity to freshwater aquatic life occurs at 

concentrations as low as 100 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity of 

a mixture of isomers of BHC to sensitive freshwater aquatic life. 

The available data for a mixture of isomers of BHC indicate 

that acute toxicity to saltwater aquatic life occurs at 

concentrations as low as 0.34 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity of 

a mixture of isomers of BHC to sensitive saltwater aquatic life. 

0 i. 
Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to hexachlorocyclohexane 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at loq5, , and 10- 

The corresponding recommended criteria are 22 ng/L, 2.2 7 

ng/L, and .22 ng/L, respectively. If these estimates are made 

for consumption of aquatic organisms only, excluding consumption 

of water, the levels are 74 ng/L, 7.4 ng/L, and .74 ng/L, 

respectively. 

10-6 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to hexachlorocyclohexane 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at and 

The corresponding recommended criteria are 134 ng/L, 13.4 ng/L, 

and 1.34 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 450 ng/L, 45.0 ng/L, and 4.50 ng/L, 

respectively. 
For the maximum protection of human health from the potential 

carcinogenic effects due to exposure of rhexachlorocyclohexane through ingestion of contaminated water and 

contaminated aquatic organisms, the ambient water concentrations 

should be zero, based on the nonthreshold assumption for this 

chemical. However, zero level may not be attainable at the 

present time. Therefore, the levels which may result in 

incremental increase of cancer risk over the lifetime are 

estimated at and The corresponding 

recommended criteria are 186 ng/L, 18.6 ng/L, and 1.86 ng/L, 

respectively. If these estimates are made for consumption of 

aquatic organisms only, excluding consumption of water, the 

levels are 625 ng/L, 62.5 ng/L, and 6.25 ng/L, respectively. 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to technicalhexachlorocyclohexane through ingestion of contaminated water 

and contaminated aquatic organisms, the ambient water 

concentrations should be zero, based on the nonthreshold 

assumption for this chemical. However, zero level may not be 

attainable at the present time. Therefore, the levels which may 

result in incremental increase of cancer risk over the 

lifetime are estimated at 10-5, loe6, and The 

corresponding recommended criteria are 52 ng/L, 5.2 ng/L, and -52 

ng/L, respectively. If these estimates are made for consumption 

of aquatic organisms only, excluding consumption of water, the 

levels are 174 ng/L, 17.4 ng/L, and 1.74 ng/L, respectively. 

, / .* , 
Using the present guidelines, satisfactory criteria cannot be 

derived at this time for d- and e- hexachlorocyclohexane because 

of insufficient available data. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
HEXACHLOROCYCLOPENTADIENE 

CRITERIA: 

Aquatic Life 

The available data for hexachlorocyclopentadiene indicate 

that acute and chronic toxicity to freshwater aquatic life occurs 

at concentrations as low as 7.0 and 5.2 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for hexachlorocyclopentadiene 

indicate that acute toxicity to saltwater aquatic life occurs 

at concentrations as low as 7.0 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. NO data are available concerning the chronic toxicity of 

0 hexachlorocyclopentadiene to sensitive saltwater aquatic life. 

Human Health 

For comparison purposes, two approaches were used to derive 

criterion levels for hexachlorocyclopentadiene. Based on 

available toxicity data, to protect public health the derived 

level is 206 ug/L. Using available organoleptic data, to 

control undersirable taste and odor quality of ambient water the 

estimated level is 1 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing 

water quality criteria, and have no demonstrated relationship to 

potential adverse human health effects. 

-. , (45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
0.3 mg/L for domestic water supplies (welfare). 

1.0 mg/L for freshwater aquatic life. 

INTRODUCTION: 

Iron is the fourth most abundant, by weight, of the elements 

that make up the earth's crust. Common in many rocks, it is an 

important component of many soils, especially the clay soils 

where usually it is a major constituent. Iron in water may be 

present in varying quantities dependent upon the geology of the 

area and other chemical components of the waterway. 

Iron is an essential trace element required by both plants 

and animals. In some waters it may be a limiting factor for the 

growth of algae and other plants; this is true especially in some 

marl lakes where it is precipitated by the highly alkaline 

conditions. It is a vital oxygen transport mechanism in the 

blood of all vertebrate and some invertebrate animals. 

The ferrous, or bivalent (Fe++), and the ferric, or trivalent 

(Fe+++) irons, are the primary forms of concern in the aquatic 

environment, although other forms may be in organic and inorganic 

wastewater streams. The ferrous (Fe++) form can persist in 

waters void of dissolved oxygen and originates usually from 

groundwaters or mines when these are pumped or drained. For 

practical purposes the ferric (Fe +++) form is insoluble. Iron 

can exist in natural organometallic or humic compounds and 

colloidal forms. Black or brown swamp waters may contain iron 

concentrations of several mg/L in the presence or absence of .. . .,_. ,~ ., .,.,. 

~~ 

dissolved oxygen, but this iron form has little effect on aquatic 
life. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 

a .. .i . :/ 
ISOPHORONE 

Aquatic Life 

The available data for isophorone indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 117,000 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of isophorone to 

sensitive freshwater aquatic life. 

The available data for isophorone indicate that acute 

toxicity to saltwater aquatic life occurs at concentrations as 

low as 12,900 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of isophorone to 

sensitive saltwater aquatic life. 

I 

Human Health 

For the protection of human health from the toxic properties 

of isophorone ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 5.2 

W/L. 

For the protection of human health from the toxic properties 

of isophorone ingested through contaminated aquatic 

organisms alone, the ambient water criterion is determined to 

be 520 mg/L. 

(45 F.R. 79318, November 28, 1980) 0 ;;. : SEE APPENDIX B FOR METHODOLOGY .. -...*. 
LEAD 

AQUATIC LIFE SUMMARY: 

The acute toxicity of lead to several species of freshwater 

animals has been shown to decrease as the hardness of water 

increases. At a hardness of 50 mg/L the acute sensitivities of 

10 species range from 142.5 ug/L for an amphipod to 235,900 ug/L 

for a midge. Data on the chronic effects of lead on freshwater 

animals are available for two fish and two invertebrate species. 

The chronic toxicity of lead also decreases as hardness increases 

and the lowest and highest available chronic values (12.26 and 

128.1 ug/L) are both for a cladoceran, but in soft and hard 

water, respectively. Acute-chronic ratios are available for 

three species and range from 18 to 62. Freshwater algae are 

affected by concentrations of lead above 500 ug/L, based on 

data for four species. Bioconcentration factors are available 

for four invertebrate and two fish species and range from 42 to 

1,700. 

0 

Acute values are available for 13 saltwater animal species 

and range from 315 ug/L for the mummichog to 27,000 ug/L for 

the soft shell clam. A chronic toxicity test was conducted 

with a mysid; unacceptable effects were observed at 37 ug/L but 

not at 17 ug/L and the acute-chronic ratio for this species is 

124.8. A species of macroalgae was affected at 20 ug/L. 

Available bioconcentration factors range from 17.5 to 2,570. 

NATIONAL CRITERIA: 0 

-2 The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 
Aquatic Organisms and Their Uses indicate that, except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the 4-day average concentration (in ug/L) of lead 

does not exceed the numerical value given by 

(1.273 [ ln(hardness) 3 -4.705) more than once every 3 years on the e 

average and if the 1-hour average concentration (in ug/L) does 

not exceed the numerical value given. by .(1.273 [ ln(hardness) 3- 

1.460) more than once every 3 years on the average. For example, 

at hardnesses of 50, 100, and 200 mg/L as CaC03 the 4-day average 

concentrations of lead are 1.3, 3.2, and 7.7 ug/L, respectively, 

and the 1-hour average concentrations are 34, 82, and 200 ug/L. 

The procedures described in, the Guidelines indicate that, 

except possibly where a locally important species is very 

sensitive, saltwater aquatic organisms and their uses should not 

be affected unacceptably if the 4-day average concentration of 

lead does not exceed 5.6 ug/L more than once every 3 years on 

the average and if the 1-hour average concentration does not 

exceed 140 ug/L more than once every three years on the average. 

EPA believes that a measurement such as tfacid-solubletl would 

provide a more scientifically correct basis upon which to 

establish criteria for metals. The criteria were developed on 

this basis. However, at this time, no EPA-approved methods for 

such a measurement are available to implement the criteria 

through the regulatory programs of the Agency and the States. 

The 

a measurement 

Agency is considering 

such as acid

development 

-soluble. 

and 

Until 

approval 

available, 

of methods 

however, 

for a 

EPA recommends applying the criteria using the total recoverable 
method. This has two impacts: (1) Certain species of some metals 

cannot be analyzed directly because the total recoverable method 

does not distinguish between individual oxidation states, and (2) 

these criteria may be overly protective when based on the total 

recoverable method. 

The recommended exceedence frequency of 3 years is the 

Agency's best scientific judgment of the average amount of time 

it will take an unstressed system to recover from a pollution 

event in which exposure to lead exceeds the criterion. A 

stressed system, for example, one in which several outfalls occur 

in a limited area, would be expected to require more time for 

recovery. The resilience of ecosystems and their ability to 

recover differ greatly, however, ' and site -specific criteria may 

be established if adequate justification is provided. 

The use of criteria in designing waste treatment facilities 

requires the selection of an appropriate wasteload allocation 

model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

impractical, in which case one should rely on a steady -state 

model. The Agency recommends the interim use of lQ5 or lQl0 for 

Criterion Maximum Concentration design flow and 745 or 7Q1Q for 

the Criterion Continuous Concentration design flow in steadystate models for unstressed and stressed systems, respectively. 

These matters are discussed in more detail in the Technical 

Support Document for Water Quality-Based Toxics Control (U.S. 

0 
HUMAN HEALTH CRITERIA: 

The ambient water quality criterion for lead is recommended 

to be identical to the existing drinking water standard which is 

50 ug/L. Analysis of the toxic effects data resulted in a 

calculated level whic is protective to human health against the 

ingestion of contaminated water and contaminated aquatic 

organisms. The calculated value is comparable to the present 

standard. For this reason a selective criterion based on 

expoeure Soley from consumption of 6.5 grams of aquatic organisms 

was not derived. 

(45 F.R. 79318 Nov. 28,1980) (50 F.R. 30784, July 29, 1985) 

SEE APPENDIX A FOR METHODOLOGY 
e CRITERION: 

MALATHION 

0.1 ug/L for freshwater and marine aquatic life. 

RATIONALE: 

The freshwater fish most sensitive to malathion, an 

organophosphorus pesticide, appear to be the salmonids and 

centrarchids. Post and Schroeder (1971) report a 96-hour LC50 

between 120 and 265 ug/L for 4 species of salmonids. Macek and 

McAllister (1970) found a 96-hour LC50 range between 101 and 285 

ug/L for 3 species of centrarchids and 3 species of salmonids. 

Other 96-hour LCSO's are: rainbow trout, -Salmo - gairdrleri, 68 ug/L 

(Cope, 1965) i largemouth bass, MicroEerus -------I salmoides 50 ug/L 

(Pickering et al. 1962); and chinook salmon, Oncorhynchus I-- 

tshawytscha ---I 23 ug/L (Katz, 1961). All of the above tests were 

in static systems. Eaton (1970) determined a 96-hour LC50 for 

bluegill, - LeEmis --------I macrochirus in a flow-through system at110 

ug/L. Macek and McAllister (1970) reported a similar 96-hOUr 

LC50 for the bluegill in a static exposure. Static 96-hour LC50s 

of 120 and 160 ug/L were reported by Post and Schroeder (1971) 

for brook trout, Salvelinus fontha&. Bender (1969) indicated 

that the acute toxicity to fathead minnows, Pimephales -- promelas, - 

is slightly greater (about 2.0 times) in a static system than in 

a f low-through system. The f low-through acute toxicity to 

fathead minnows reported by Mount and Stephan (1967) approximated 

the static acute toxicity reported by Henderson and Pickering 

0 

(1958) and Bender (1969). 0 

xi 
Many aquatic invertebrates appear to be more sensitive than 

fish to malathion. The 96-hour LC50 for Gammarus lacustris was 

1.0 ug/L (Sanders, 1969); for Pteronarcella badia 1 1.1 ug/L 

(Sanders and Cope, 1968); and for ---- Gammarus ----- fasciatus 1 0.76 ug/L 

(Sanders, 1972). The 48-hOUr LC50 for Simocephalus -----serrulatus 

was 3.5 ug/L and for - Daphnia - pulex, 1.8 ug/L (Sanders and Cope, 

1966). -- Daphnia ---- were immobilized in 50 hours in 0.9 ug/L 

(Anderson, 1960). The 24-hour LC50s for two species of midge 

larvae were 2.1 ug/L (Mulla and Xhasawinah, 1969) and 2.0 ug/L 

(Karnak and Collins, 1974). 

Safe life cycle exposure concentrations for the more 

sensitive invertebrates are not known. The most sensitive 

aquatic organisms probably have not yet been tested; safe 

concentrations for the most s'ensitive invertebrates exposed 

through a complete life cycle have not been determined; and 

effects of low concentrations on invertebrate behavior are 

unknown. 

The stability of malathion in water is dependent on the 

chemical and biological conditions of the water (Paris et al. 

1975). Weiss and Gakstatter (1964) have shown that the half-life 

of malathion was reduced from about 5 months at pH 6 to 1 to 2 

weeks at pH 8. Eichelberger and Lichtenberg (1971) found that 

only 10 percent remained in the Little Miami River (pH 7.3-8.0) 

after 2 weeks. Bender (1969) states that one of the malathion 

breakdown products may be more toxic than the parent compound. 

It has been shown that a measured concentration of 575 ug/L 

malathion in flowing seawater kills 40 to 60 percent of the 
marine fish, Lagodon rhomboides, in 3.5 hours and causes about 75 

percent brain acetylcholinesterase (AChE) inhibition (Coppage et 

al. 1975). Similar inhibition of AChE and mortality were caused 

in pinfish in 24, 48, and 72 hours'at measured concentrations of 

142, 92 and 58 ug/L, respectively. A concentration of 31 ug/L 

caused 34 percent AChE inhibition in pinfish but no deaths in 72 

hours. Coppage and Matthews (1974) demonstrated that death may 

be associated with reductions of brain AChE activity of four 

marine fishes by 70 to 80 percent or more in short-term exposures 

to malathion. Coppage and Duke (1971) found that moribund 

mullet, Mugil cephalus, in an estuary sprayed with malathion (3 

oz./acre) during a large-scale mosquito control operation had 

about 98 percent inhibition of brain AChE. This is in agreement 

with 70 to 80 percent or more inhibition of brain AChE levels at 

and below which some deaths are likely to occur in short-term 

exposure. Spot, Leiostomus xanthurus, and Atlantic croaker, 

Micropogon - undule, also had substantial inhibition of brain 

during the spray operation (70 percent or more inhibition). 

0 

Toxicity studies have been made on a number of marine 

animals. Eisler (1970) studied the 96-hour LC50 for several 

marine fishes at 20 OC in static, aerated seawater. The 96-hour 

LC50 values (in ug/L) were: Menidia menidia, 125: - Mugil cephalus, 

550; Fundulus rnnialis _--- 1 250; Fundulus heteroclitus, 240; 

- sphaeroides --_-__--- ---_-----I maculatus 3250: __ Anquilla _-___ --___-__I rostrata 82: and 

Thalassoma bifasciatum, 27. Katz (1961) reported the static 24- 

hour LC50 for Gasterosteus aculeatus in 25 o/oo saltwater as 76.9 

ug/L active ingredient. The 96-hour LC50 for striped bass, 
_--_ Morone --.-------I saxatilis in intermittent flowing seawater has been 

reported as 14 ug/L (U.S. BSFW, 1970). 

Reporting on studies of the toxicity of malathion on marine 

invertebrates, Eisler (1969) found the 96-hour LC50 (static, 24 

o/oo salinity aerated) to be 33 ug/L for sand shrimp, Crangon 

septemspinosa; 82 ug/L for grass shrimp, Saemonetes vulgaris; -- 

and 83 ug/L for hermit crab, Pagurus longicarpus. Growth of 

oyster, Crassostrea virginica, was reduced 32 percent by 96-hour 

exposure to 1 mg/L (Butler, 1963). The 48- hour LC50 for 

fertilized eggs of oysters was estimated by Davis and Hidu (1969) 

to be 9.07 mg/L and the 14-day LC50 for larvae, 2.66 mg/L. 

Malathion enters the aquatic environment primarily as a 

result of its application as an insecticide. Because it degrades 

quite rapidly in most waters, depending on pH, its occurrence is 

sporadic rather than continuous. Because the toxicity is exerted 

through inhibition of AChE and because such inhibition may be 

additive with repeated exposures and may be caused by any of the 

organophosphorus insecticides, inhibition of AChE by more than 35 

percent may be expected to result in damage to aquatic organisms. 

An application factor of 0.1 is applied to the 96-hOUr LC50 

data for Gammarus lacustris, - G. fasciatis and Daphnia, which are 

all approximately 1.0 ug/L, yielding a criterion of 0.1 ug/L. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
MANGANESE 

50 ug/L for domestic water supplies (welfare): 

100 ug/L for protection of consumers of marine molluscs. 

INTRODUCTION: 

Manganese does not occur naturally as a metal but is found in 

various salts and minerals, frequently in association with iron 

compounds. The principal manganese-containing substances are 

manganese dioxide (MnOZ) , pyrolusite, manganese carbonate 

(rhodocrosite) and manganese silicate (rhodonite) ~ The oxides 

are the only important minerals mined. Manganese is not mined in 

the United States except when manganese is contained in iron ores 

that are deliberately used to foh ferro-manganese alloys. 

The primary uses of manganese are in metal alloys, dry cell 

batteries, micro-nutrient fertilizer additives, organic compounds 

used in paint driers and as chemical reagents. Permanganates are 

very strong oxidizing agents of organic materials. 

Manganese is a vital micro-nutrient for both plants and 

animals. When manganese is not present in sufficient quantities, 

plants exhibit chlorosis (a yellowing of the leaves) or failure 

of the leaves to develop properly. Inadequate quantities of 

manganese in domestic animal food results in reduced reproductive 

capabilities and deformed or poorly maturing young. Livestock 

feeds usually 

high corn diet 

have sufficient manganese, but beef cattle on a 

may require a supplement. 
RATIONALE: 

Although inhaled manganese dusts have been reported to be 

toxic to humans, manganese normally is ingested as a trace 

nutrient in food. The average human intake is approximately 10 

mg/day (Sollman, 1957). Very large doses of ingested manganese 

can cause some disease and liver damage but these are not known 

to occur in the United States. Only a few manganese toxicity 

problems have been found throughout the world and these have 

occurred under unique circumstances, i.e., a well in Japan near a 

deposit of buried batteries (McKee and Wolf, 1963). 

It is possible to partially sequester manganese with special 

treatment but manganese is not removed in the conventional 

treatment of domestic waters (Riddick et al. 1958: Illig, 1960). 

Consumer complaints arise when manganese exceeds a concentration 

of 150 ug/L in water supplies (Griffin, 1960). These complaints 

are concerned primarily with the brownish staining of laundry and 

objectionable tastes in beverages. It is possible that the 

presence of low concentrations of iron may intensify the adverse 

effects of manganese. Manganese at concentrations of about 10 to 

20 ug/L is acceptable to most consumers. A criterion for 

domestic water supplies of 50 ug/L should minimize the 

objectionable qua1 ities. 

McKee and Wolf (1963) summarized data on toxicity of 

manganese to freshwater aquatic life. Ions of manganese are 

found rarely at concentrations above 1 mg/L. The tolerance 

values reported range from 1.5 mg/L to over 1000 mg/L. Thus, 

manganese is not considered to be a problem in fresh waters. 

Permanganates havebeen reportedtokill fish in 8to 18 hours at 
concentrations of 2.2 to 4.1 mg/L, but permanganates are not 

persistent because they rapidly oxidize organic materials and are 

thereby reduced and rendered nontoxic. 

Few data are available on the toxicity of manganese to marine 

organisms. The ambient concentration of manganese is about 2 ug/L 

(Fairbridge, 1966). The material is rapidly assimilated and 

bioconcentrated into nodules that are deposited on the sea floor. 

The major problem with manganese may be concentration in the 

edible portions of molluscs, as bioaccumulation factors as high 

as 12,000 have been reported (NAS, 1974). In order to protect 

against a possible health hazard to humans by manganese 

accumulation in shellfish, a criterion of 100 ug/L is recommended 

for marine water. 

Manganese is not known to be a problem iq water consumed by 

livestock. At concentrations of slightly less than 1 mg/L to a 

few milligrams per liter, manganese nay be toxic to plants from 

irrigation water applied to soils with pH values lower than 6.0. 

The problem may be rectified by liming soils to increase the pH. 

Problems may develop with long-term (20 year) continuous 

irrigation on other soils with water containing about 10 mg/L of 

manganese (NAS, 1974). But, as stated above, manganese is rarely 

found in surface waters at concentrations greater than 1 mg/L. 

Thus, no specific criterion for manganese in agricultural waters 

is proposed. In select areas, and where acidophilic crops ate 

cultivated and irrigated, a criterion of 200 ug/L is suggested 0 for consideration. i \+./ 
Most industrial users of water can operate successfully where 

the criterion proposed for public water supplies is observed. 

Examples of industrial tolerance of manganese in water are 

summarized for industries such as dyeing, milk processing, paper, 

textiles, photography and plastics (McKee and Wolf, 1963). A more 

restrictive criterion may be needed to protect or ensure product 

quality. 

(QUALITY CRITERIA FOR WATER, JULY 1976) P B - 263943 

SEE APPENDIX C FOR METHODOLOGY 
*MERCURY 

AQUATIC LIFE 

Data are 

SUMMARY : 

available on the acute toxicity of mercury(I1) ,to 28 

genera of freshwater animals. Acute values for invertebrate 

species range from 2.2 ug/L for Daphnia pulex to 2,000 ug/L for 

three insects. Acute values for fishes range from 30 ug/L for 

the guppy to 1,000 ug/L for the - Mozambique tilapia. Few data are 

available for various organomercury compounds and mercurous 

nitrate, and they all appear to be 4 to 31 times more acutely 

toxic than mercury(I1). 

Available chronic data indicate that methylmercury is the 

most chronically toxic of the tested mercury compounds. Tests on 

methylmercury with Daphnia magna and brook trout produced chronic 

values less than 0.07 ug/L. For mercury(I1) the chronic value 

obtained with - Daania __ - msna - was about 1.1 ug/L and the acute - 

chronic ratio was 4.5. In both a life - cycle test and an early 

life -stage test on mercuric chloride with the fathead minnow, the 

chronic value was less than 0.26 ug/L and the acute-chronic ratio 

was over 600. 

0 

Freshwater plants show a wide range of sensitivities to 

mercury, but the most sensitive plants appear to be less 

sensitive than the most sensitive freshwater animals to both 

mercury(I1) and methylmercury. A bioconcentration factor of 

4,994 is available for mercury(II), but the bioconcentration 

factors for methylmercury range from 4,000 to 85,000. 0 

\., *Indicates suspended, canceled or restricted by U.S. EPA 

Office of Pesticides and Toxic Substances 
Data on the acute toxicity of mercuric chloride are available 

for 29 genera of saltwater animals, including annelids, molluscs, 

crustaceans, echinoderms, and fishes. Acute values range from 

3.5 ug/L for a mysid to 1,678 ug/L for winter flounder. Fishes 

tend to be more resistant and molluscs and crustaceans tend to be 

more sensitive to the acute toxic effects of mercury(I1). 

Results of a life-cycle test with the mysid show that mercury(I1) 

at a concentration of 1.6 ug/L significantly affected time of 

first spawn and productivity; the resulting acute-chronic ratio 

was 3.1. 

Concentrations of mercury that affected growth and 

photosynthetic activity of one saltwater diatom and six species 

of brown algae range from 10 to 160 ug/L. Bioconcentration 

factors of 10,000 and 40,000 have been obtained for mercuric 

chloride and methylmercury with an oyster. 

NATIONAL CRITERIA: 

Derivation of a water quality criterion for mercury is more 

complex than for most metals because of methylation of mercury in 

sediment, in fish, and in the food chain of fish. Apparently 

almost all mercury currently being discharged is mercury(I1). 

Thus mercury(I1) should be the only important possible cause of 

acute toxicity and the Criterion Maximum Concentrations can be 

based on the acute values for mercury(I1). 

The best available data concerning long-term exposure of fish 

to mercury(I1) indicates that concentrations above 0.23 ug/L 

caused statistically significant effects on the fathead minnow 

and caused the concentration of total mercury in the whole body 
to exceed 1.0 mg/kg. Although it is not known what percent of 

the mercury in the fish was methylmercury, it is also not known 

whether uptake from food would increase the concentration in the 

fish in natural situations. Species such as rainbow trout, coho 

salmon, and especially the bluegill, might suffer chronic effects 

and accumulate high residues of mercury about the same as the 

fathead minnow. 

With regard to long-term exposure to methylmercury, McKim et 

al. (1976) found that brook trout can exceed the FDA action level 

without suffering statistically significant adverse effects on 

survival, growth, or reproduction. Thus for methylmercury the 

Final Residue Value would be substantially lower than the Final 

Chronic Value. 

Basing a freshwater criterion on the Final Residue Value of 

0.012 ug/L derived from the bioconcentration factor of 81,700 for 

methylmercury with the fathead minnow (Olson et al. 1975) 

essentially assumes that all discharged mercury is methylmercury. 

On the other hand, there is the possibility that in field 

situations uptake from food might add to the uptake from water. 

Similar considerations apply to the derivation of the saltwater 

criterion of 0.025 ug/L using the BCF of 40,000 obtained for 

methylmercury with the Eastern oyster (Kopfler, 1974). Because 

the Final Residue Values for methylmercury are substantially 

below the Final Chronic Values for mercury(II), it is probably 

not too important that many fishes, including the rainbow trout, 

coho salmon, bluegill, and haddock might not be adequately 

protected by the freshwater and saltwater Final Chronic Values 

for mercury(I1). 
In contrast to all the complexities of deriving numerical 

criteria for mercury, monitoring for unacceptable environmental 

effects should be relatively straightforward. The most sensitive 

adverse effect will probably be exceedence of the FDA action 

level. Therefore, existing discharges should be acceptable if 

the concentration of methylmercury in the edible portion of 

exposed consumed species does not exceed the FDA action level. 

The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses indicate that, except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the 4-day average concentration of mercury does I 

not exceed 0.012 ug/L more than once every 3 years on the average 

and if the 1-hour average concentration does not exceed 2.4 ug/L 

more than once every 3 years on the average. If the 4-day 

average concentration exceeds 0.012 ug/L more than once in a 3- 

year period, the edible portion of consumed species should be 

analyzed to determine whether the concentration of methylmercury 

exceeds the FDA action level. 

The procedures described in the Guidelines indicate that, 

except possibly where a localy important species is very 

sensitive, saltwater aquatic organisms and their uses should not 

be affected unacceptably if the 4-day average concentration of 

mercury does not exceed 0.025 ug/L more than once every 3 years 

on the average and if the 1-hour average concentration does not 

exceed 2.1 ug/L more than once every 3 years on the average. If 

the 4-day average concentration exceeds 0.025 ug/L more than once 
in a 3-year period, the edible protion of consumed species should 

be analyzed to determine whether the concentration of 

mathylmercury exceeds the FDA action level. 

0 

EPA believes that a measurement such as "acid-soluble" would 

provide a more scientifically correct basis upon which to 

establish criteria for metals. The criteria were developed on 

this basis. However, at this time, no EPA approved-methods for 

such a measurement are available to implement the criteria 

through the regulatory programs of the Agency and the States. 

The Agency is considering development and approval of methods for 

a measurement such as acid-soluble. Until available, however, 

EPA recommends applying the criteria using the total recoverable 

method. This has two impacts: (1) certain species of some metals 

cannot be analyzed directly because the total recoverable method 

does not distinguish between individual oxidation states, and (2) 

these criteria may be overly protective when based on the total 

recoverable method. 

0 

The recommended exceedence frequency of 3 years is the 

Agency's best scientific judgment of the average amount of time 

it will take an unstressed system to recover from a pollution 

event in which exposure to mercury exceeds the criterion. A 

stressed system, for example, one in which several outfalls occur 

in a limited area, would be expected to require more time for 

recovery. The resilience of ecosystems and their ability to 

recover differ greatly, however, and site -specif ic criteria may 

0 be established if adequate justification is provided. .., ,' 
The use of criteria in designing waste treatment facilities 

requires the selection of an appropriate wasteload a1 location 

model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

impractical, in which case one should rely on a steady-state 

model. The Agency recommends the interim use of 1Q5 or lQl0 for 

Criterion Maximum Concentration design flow and 745 or 7410 for 

the Criterion Continuous Concentration design flow in steadystate models for unstressed and stressed systems respectively. 

These matters are discussed in more detail in the Technical 

Support Document for Water Quality-Based Toxics Control (U.S 

EPA, 1985). 

HUMAN HEALTH CRITERIA 

I 

For the protection of human health from the toxic properties 

of mercury ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 144 

nWL. 

For the protection of human health from the toxic properties 

of mercury ingested through contaminated aquatic organisms alone, 

the ambient water criterion is determined to be 146 ng/L. 

NOTE: These values include the consumption of freshwater, 

estuarine, and marine species. 

(45 F.R. 79318 Nov. 28,1980) (50 F.R. 30784, July 29, 1985) 

SEE APPENDIX A FOR METHODOLOGY 
0 CRITERIA: 

METHOXYCHMR 

100 ug/L for domestic water supply (health); 

0.03 ug/L for freshwater and marine aquatic life. 

RATIONALE : 

The highest level of methoxychlor found to have minimal or no 

long-term effects in man is 2.0 mg/kg of body weight/day (Lehman, 

1965). Where adequate human data are available for corroboration 

of the animal results, the total tfsafe't drinking water intake 

level is assumed to be l/lOO of the no-effect or minimal effect 

level reported for the most sensitive animal tested, in this 

case, man. 

Applying the available data and based upon the assumptions 

that 20 percent of the total intake of methoxychlor is from 

drinking water, and that the average person weighs 70 kg and 

consumes 2 liters of water per day, the formula for calculating a 

criterion is 2.0 mg/kg x 0.2 x 70 kg x 1/100 x 1/2 = 0.14 mg/L. 

A criterion level for domestic water supply of 100 ug/L is 

recommended. 

Few data: are available on acute and chronic effects of 

methoxychlor on freshwater fish. Merna and Eisele (1973) 

observed reduced hatchability of fathead minnow pimzhales - 

prcelas) -- embryos at 0.125 ug/L and lack of spawning at 2.0 

ug/L. Yellow perch, Perca -- flavescens, exposed to 0.6 ug/L for 8 

months exhibited reduced growth. The 36-hour LC50 concentration 

was 7.5 and 22 ug/L for the fathead minnow and yellow perch, 

.- respectively. Korn and Earnest (1374) obtained a 96-hour LC50 
of 3.3 ug/L with juvenile stripped bass, Morone -----_I saxatilis 

exposed to methoxychlor in a flowing-water bioassay. 

Sanders (1972) determined a 96-hour LC50 value of 0.5 ug/L 

for the crayfish, Orconectes nais. Merna and Eisele (1973) 

obtained a 96-hour LC50 value of 0.61 ug/L for the scud, Gammarus 

pseudolimnaeus and 96-hour LC5O's ranging from 1.59 to 7.05 ug/L 

for the crayfish, Orconectes --I nais and three aquatic insect 

larvae. In 28-day exposuresI reduction in emergence of mayflies, 

----- Stenonema sp. and in pupation of caddisf lies, Cheumatospsyche _- 

sp., were observed at 0.5 and 0.25 ug/L concentrations, 

respectively. They also found methoxychlor to be degraded in a 

few weeks or less in natural waters. 

Eisele (1974) conducted a study in which a section of a 

natural stream was dosed at 0.2 ug/L methoxychlor for 1 year. 

The near extinction of one species of scud, wllella ---- azteca, and 

reductions in populations of other sensitive species, as well as 

biomass, were observed. Residue accumulation of up to 1,000 

times the level in the stream was observed in first-year 

crayfish, Orconectes __ nais. Metcalf et al. (1971) traced the 

rapid conversion of methoxychlor to water soluble compounds and 

elimination from the tissues of snails, mosquito larvae and 

mosquitofish. Thus, methoxychlor appears to be considerably less 

bioaccumlative in aquatic organisms than some of the other 

chlorinated pesticides. 

Methoxyhlor has a very low accumulation rate in birds and 

mammals (Stickel, 1973), and relatively low avian (Heath et al. 

1972) and mammalian (Hodge et al. 1950) toxicities. No 

administrative guidelines for acceptable levels in edible fish 
tissues have been established by the U.S. Food and Drug 

Administration. 

The above data indicate that 0.1 ug/L methoxychlor would be 

just below chronic effect level for the fathead minnow and onefifth the acute toxicity level in a crayfish species. Therefore, 

a criterion level of 0.03 ug/L is recommended. This criterion 

should protect fish as sensitive as striped bass and is 10 times 

lower than the level causing effects on some invertebrate 

populations in a 1-year dosing of a natural stream. 

Bahner and Nimmo (1974) found the 96-hour LC50 of 

methoxychlor for the pink shrimp, Penaeus ----I duorarum to be 3.5 

ug/L and the 30-day LC50 to be 1.3 ug/L. Using an application 

factor of 0.01 with the pink shrimp's acute toxicity of 3.5 

ug/L, the recommended criterion for the marine environment is 

0.03 ug/L. 

Butler (1971) found accumulation factors of 470 and 1,500 for 

the molluscs, _-_____-_- Mercenaria ______-___ mercenaria and gyg arenaria, 

respectively, when exposed to 1 ug/L methoxychlor for 5 days. 

Using the 1,500 accumulation factor as a basis, a water 

concentration of 0.2 ug/L would be required to meet the U.S. Food 

and Drug Administration's guideline for methoxychlor in meat 

products. Thus, the recommended marine criterion of 0.03 ug/L is 

an order of magnitude lower than this concentration. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY - ., 
MIREX 

CRITERION: 

0.001 ug/L for freshwater and marine aquatic life. 

RATIONAL??.: 

Mirex is used to control the imported fire ant Solenzsis - 

saevissima richteri in the southeastern United States. Its use 

is essentially limited to the control of this insect and it is 

always presented in bait. In the most common formulation, 

technical grade mirex is dissolved in soybean oil and sprayed on 

corncob grits. The bait produced in this manner consists of 0.3 

percent mirex, 14.1 percent soybean oil and 85 percent corncob 

grits. The mirex bait often is applied at a rate of 1.4 kg/ha, 

I 

equivalent to 4.2 grams of toxicant per hectare. 

Relatively few studies have been made of the effects of mirex 

on freshwater invertebrates og yhrdr, only Dudke et al. (1971) 

report chemical analyses of mirex in the water. Their study 

reported effects on two crayfish species exposed to mirex by 

three techniques. First, f ield-col lected crayfish were exposed 

to several sublethal concentrations of technical grade mirex 

solutions €or various periods of time; second, crayfish were 

exposed to mirex leached from bait (0.3 percent active 

ingredient) ; and third, the crayfish were fed mirex bait. 

Procambarus uandingi juveniles were exposed to 1 or 5 ug/L 

for 6 to 144 hours, transferred to clean water and observed for 

10 days. After 5 days in clean water, 95 percent of the animals 

exposed to 1 ug/L for 14 hours were dead. Exposure to 5 ug/L for 

6, 24, and 58 hours resulted in 26, 50, and 98 percent mortality 

10 days after transfer to clean water. Crayfish, Procambarus 

0 
’.. 

- hayi, were exposed to 0.1 and 0.5 ug/L for 48 hours. Four days 

after transfer to clean water, 65 percent of the animals exposed 

to 0.1 ug/L were dead. At the 0.5 ug/L concentration, 71 

percent of the animals were dead after 4 days in clean water. 

Tissue residue accumulations (wet weight basis) ranged from 940- 

to 27,210-fold above water concentrations. In leached bait 

experiments, 10 bait particles were placed in 2 liters of water 

but isolated from 20 juvenile crayfish. Thirty percent of the 

crayfish were dead in 4 days and 95 percent were dead in 7 days. 

Water analysis indicated mirex concentrations of 0.86 ug/~. In 

feeding Qxperiments, 108 crayfish each were fed one bait 

particle. Mortality was noticed on the first day after feeding, 

j8 and by the sixth day 77 percent were dead. In another experiment, 

all crayfish were dead 4 days after having been fed 2 bait 

particles each. From this report it is obvious that mirex is 

extremely toxic to these species of crayfish. Mortality and 

accumulation increases with time of exposure to the insecticide. 

Concentrations as low as 0.1 ug/L or the ingestion of one 

particle resulted in death. 

$ 

Research to determine effects of mirex on fish has been 

concentrated on species which have economic and sport fishery 

importance. Hyde et al. (1974) applied mirex bait (0.3 percent 

mirex) at the standard rate (1.4 kg/ha) in four ponds containing 

channel catfish, Ictalurus punctatus. Three applications were 

made over an 8~011th period with the first application 8 days 

after fingerling (average weight 18.4 g) catfish were placed in 

the ponds. Fish were collected at each subsequent application 
(approximately 4-month intervals). Two and one half months after 

the final application, the ponds were drained, all fish were 

measured and weighed, and the percent survival was calculated. 

Mirex residues in the fish at termination of the experiment 

ranged from 0.015 ug/g (ppm) in the fillet to 0.255 uq/g in the 

fat. 

0 

In another study, Van Valin et al. (1988) exposed bluegills, 

- Lepomis macrochirus, and the goldfish, Carassius auratus, to 

mirex by feeding a mirex-treated diet (1, 3, and 5 mg mirex per 

kg body weight) or by treating holding ponds with mirex bait 

(1.3, 100, and 1000 ug/L computed water concentration). They 

reported no mortality or tissue pathology for the bluegills; 

however, after 58 days of exposure, gill breakdown in goldfish 

was found in the 100 and 1000 ug/L contact exposure ponds, and 

kidney breakdown was occurring in the 1000 ug/L ponds. Mortality 

in the feeding experiments was not related to the level of 

exposure, although growth of the bluegills fed 5 ug/L mirex was 

reduced. 

In laboratory and field test systems, reported concentrations 

of mirex usually are between 0.5 and 1.0 ug/L (Van Valin et al. 

1968: Ludke et al. 1971). Although mirex seldom is found above 1 

ug/L in the aquatic environment, several field studies have shown 

that the insecticide is accumulated through the food chain. 

Borthwick et al. (1973) reported the accumulation of mirex in 

South Carolina estuaries. Their data revealed that mirex was 

transported from treated land and marsh to the estuary animals 

and that accumulation, especially in predators, occurred. In the 

test area, water supplies consistently were less than 0.01 ug/L. 
Residues in fish varied from non - detectable 

percent of the samples containing 

and the percent of samples containing mirex 

residues. 

to 0.8 ug/g with 15 

The amount of mirex 

increased at higher 

trophic levels. Fifty -four percent of the raccoons sampled 

contained mirex residues up to 4.4 ug/g and 78 percent of the 

birds contained residues up to 17 ug/g. Navgi and de la Cruz 

(1973) reported average residues for molluscs (0.15 ug/g), fish 

(0.26 ug/g), insects (0.29 ug/g), crustaceans (0.44 ug/g) and 

annelids (0.63 ug/g. They also reported that mirex was found in 

areas not treated with mirex which suggests movement of the 

pesticide in the environment. Wol fe and Norment (1973) sampled 

an area for one year following an aerial application of mirex 

bait (2.1 g mirex/ha). Crayfish residues ranged from 0.04 to 

0.16 ug/g. Fish residues were about 2 to 20 times greater than 

the controls and averaged from 0.01 to 0.78 ug/g. Kaiser (1974), 

reported the presence of rnirex in fish from the Bay of Quinte, 

Lake Ontario, Canada. Concentrations range from 0.02 ug/g in 

the gonads of the northern long nose gar, Lepistosteus osseus, to 

0.05 ug/g in the areal fin of the northern pike, ___- Esox lucius. 

Mirex has never been registered for use in Canada. 

Mirex does not appear to be greatly toxic to birds, with 

LCSO’s for the young of four species ranging from 547 to greater 

than 1667 ug/g (Heath et al. 1972). Long-term dietary dosages 

caused no adverse effect at3 ug/gwithmallards and13 ug/gwith 

pheasants (Heath and Spann, 1973). However, it has been reported 

(Stoke et al. 1978) that the persistence of mirex in bird tissue 

exceeds that of all organochlorine compounds tested except for 
DDE. Delayed mortality occurred among birds subjected to doses 

above expected environmental concentration. 

A summary examination of the data available at this time 

shows a mosaic of effects. Crayfish and channel catfish survival 

is affected by mirex in the water or by ingestion of the bait 

particles. Bioaccumulation is well established for a wide 

variety of organisms but the effect of this bioaccumulation on 

the aquatic ecosystem is unknown. There is evidence that mirex 

is very persistent in bird tissue. Considering the extreme 

toxicity and potential for bioaccumulation, every effort should 

be made to keep mirex bait particles out of water containing 

aquatic organisms and water concentrations should not exceed 

0.001 ug/L mirex. This value is based upon an application factor 

of 0.01 applied to the lowest levels at which effects on crayfish 

have been observed. 

Data upon which to base a marine criterion involve several 

estuarine and marine crustaceans. A concentration of 0.1 ug/L 

technical grade mirex in flowing seawater was lethal to juvenile 

pink shrimp, Penaeus durorarum, in a 3-week exposure (Lowe et al. 

1971). In static tests with larval stages (megalopal) of the mud 

crab, Rhithropanopeus harrisii, reduced survival was observed in 

0.1 ug/L mirex (Bookhout et al. 1972). In three of four 28-day 

seasonal f low-through experiments, Tagatz et al. (1975) found 

reduced survival of __Callinectes - sapidus, Penaeus durorarum, and 

grass shrimp, -_-_--_--_ Palaemonetes ~us&, at levels of 0.12 ug/L in 

summer, 0.06 ug/L in fall and 0.09 ug/L in winter. 

Since two reports, Lowe et al. (1971) and Bookhout et al. - 

(1972), stated that effects of mirex on estuarine and marine 
crustaceans were observed only after considerable time had 

elapsed, it seems reasonable that length of exposure is an 

important consideration €or this chemical. This may not be the 

case in fresh water since the crayfish were affected within 48 

hours. Therefore, a 3- to 4- week exposure might be considered 

88acute81 and by applying an application factor of 0.01 to a 

reasonable average of toxic - effect levels as summarized above, a 

recommended marine criterion of 0.001 ug/L results. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
N?QrnLENE 

Aquatic Life 

The available data for naphthalene indicate that acute and 

chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 2,300 and 620 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for naphthalene indicate that acute 

toxicity to saltwater aquatic 1 ife occurs at concentrations as 

low as 2,350 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of naphthalene to 

sensitive saltwater aquatic life. 

I 

Human Health 

Using the present guidelines, a satisfactory criterion cannot 

be derived at this time because of insufficient available data 

for naphthalene. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
NICKEL 

Aquatic Life 

FOI total recoverable nickel the criterion (in ug/L) to 

protect freshwater aquatic life as derived using the Guidelines 

is the numerical value given by e(0.76[1n(hardness)]+l.06) as a 

24-hour average, and the concentration (in ug/L) should not 

exceed the numerical value given by .(0.76[ln (hardness)]+4.02) 

at any time. For example, at hardnesses of 50, 100, and 200 mg/L 

as CaC03 the criteria are 56, 96, and 160 ug/L, respectively, as 

24-hour averages, and the concentrations should not,exceed 1,100, 

1,800, and 3,100 ug/L, respectively, at any time. 

For total recoverable nickel the criterion to protect 

saltwater aquatic life as derived using the Guidelines is 7.1 

ug/L as a 24-hour average, and the concentration should not 

exceed 140 ug/L at any time. 

0 

Human Health 

For the protection of human health from the toxic properties 

of nickel ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 632 

wl/L. 

For the protection of human health from the toxic properties 

of nickel ingested through contaminated aquatic organisms 

alone, the ambient water criterion is determined to be 4.77 

W/L. 0 (45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
0 CRITERION: 

NITF?ATES/NITRITES 

10 mg/L nitrate nitrogen (N) for 

domestic water supply (health). 

INTRODUCTION: 

Two gases (molecular nitrogen and nitrous oxide) and five 

forms of nongaseous, combined nitrogen (amino and amide groups, 

ammonium, nitrite, and nitrate) are important in the nitrogen 

cycle. The amino and amide groups are found in soil organic 

matter and as constituents of plant and animal protein. The 

ammonium ion either is released from proteinaceous organic matter 

and urea, or is synthesized in industrial processes involving 

atmospheric nitrogen fixation. The nitrite ion is formed from 

the nitrate or the ammonium ions by certain microorganisms found 

in soil, water, sewage, and the digestive tract. The nitrate ion 

is formed by the complete oxidation of ammonium ions by soil or 

water microorganisms; nitrite is an intermediate product of this 

nitrification process. In oxygenated natural water systems 

nitrite is rapidly oxidized to nitrate. Growing plants 

assimilate nitrate or ammonium ions and convert them to protein. 

A process known as denitrification takes place when nitratecontaining soils become anaerobic and the conversion to nitrite, 

molecular nitrogen, or nitrous oxide occurs. Ammonium ions may 

also be produced in some circumstances. 

0 

Among the major point sources of nitrogen entry into water 

bodies are municipal and industrial wastewaters, septic tanks, 

and feed lot discharges. Diffuse sources of nitrogen include a 

farm-site fertilizer and animal wastes, lawn fertilizer, leachate 
from waste disposal in dumps or sanitary landfills, atmospheric 

fallout, nitric oxide and nitrite discharges from automobile 

exhausts and other combustion processes, and losses from natural 

sources such as mineralization of soil organic matter (NAS, 

1972). Water reuse systems in some fish hatcheries employ a 

nitrification process for ammonia reduction; this may result in 

exposure of the hatchery fish to elevated levels of nitrite 

(Russo et al. 1974). 

RATIONALE : 

In quantities normally found in food or feed, nitrates become 

toxic only under conditions in which they are, or may be, reduced 

to nitrites. Otherwise, at "reasonable" concentration nitrates 

are rapidly excreted in the urine. High intake of nitrates 

constitutes a hazard primarily to warmblooded animals under 

conditions that are favorable to reduction to nitrite. Under 

certain circumstances, nitrate can be reduced to nitrite in the 

gastrointestinal tract which then reaches the bloodstream and 

reacts directly with hemoglobin to produce methemoglobin, 

consequently impairing transport. 

I 

The reaction of nitrite with hemoglobin can be hazardous in 

infants under 3 months of age. Serious and occasionally fatal 

poisonings in infants have occurred following ingestion of 

untreated well waters shown to contain nitrate at concentrations 

greater than 10 mg/L nitrate nitrogen (N) (NAS, 1974). High 

nitrate concentrations frequently are found in shallow farm and 

rural community wells, often as the result of inadequate 

protection from barnyard drainage or from septic tanks (USPHS, 
1961; Stewart et al. 1967). Increased concentrations of nitrates 

also have been found in streams from farm tile drainage in areas 

of intense fertilization and farm crop production (Harmeson et 

al. 1971). Approximately 2,000 cases of infant methemoglobinemia 

have been reported in Europe and North America since 1945; 7 to 

8 percent of the affected infants died (Walton, 1951; 

Sattelmacher, 1962). Many infants have drunk water in which the 

nitrate nitrogen content was greater than 10 mg/L without 

developing methemoglobinemia. Many pub1 ic water supplies in the 

United States contain levels that routinely exceed this amount, 

but only one U.S. case of infant methemoglobinemia associated 

with a public water supply has ever been reported (Virgil et al. 

1965). The differences in susceptibility to methemoglobinemia 

are not yet understood but appear to be related to a combination 

of factors including nitrate concentration, enteric bacteria, and 

the lower acidity characteristic of the digestive systems of baby 

mammals. Methemoglobinemia systems and other toxic effects were 

observed when high nitrate well waters containing pathogenic 

bacteria were fed to laboratory mammals (Wolff et al. 1972). 

Conventional water treatment has no significant effect on nitrate 

removal from water (NAS, 1974). 

Because of the potential risk of methemoglobinemia to bottlefed infants, and in view of the absence of substantiated 

physiological effects at nitrate concentrations below 10 mg/L 

nitrate nitrogen, this level is the criterion for domestic water 

supplies. Waters with nitrite nitrogen concentrations over 1 0 ~ 
mg/L should not be used for infant feeding. Waters with a 

significant nitrite concentration usually would be heavily 

polluted and probably bacteriologically unacceptable. 

Westin (1974) determined that the respective 96-hOUr and 7- 

day LC50 values for chinook salmon, Oncorhynchus tshawytscha, 

were 1,310 and 1,080 mg/L nitrate nitrogen in fresh water and 

990 and 900 mg/L nitrate nitrogen in 15 o/oo saline water. For 

fingerling rainbow trout, Salmo qairdneri, the respective 96-hour 

and 7-day LC50 values were 1,360 and 1,060 mg/L nitrate nitrogen 

in fresh water, and 1,050 and 900 mg/L nitrate nitrogen in 15 

o/oo saline water. Trama (1954) reported that the 96-hOUr LC50 

for bluegills, Asomis -- macrochirus ------I at 20°C was 2,000 mg/L 

nitrate nitrogen (sodium nitrate) and 420 mg/L nitrate nitrogen 

(potassium nitrate). Knepp and Arkin (1973) observed that 

largemouth bass, Micropterus salmoides, and channel catfish, 

Ictalurus punctatus, could be maintained at concentrations up to 

400 mg/L nitrate (90 mg/L nitrate nitrogen) without significant 

effect upon their growth and feeding activities. 

The 96-hour and 7-day LC50 values for chinook salmon, 

-___ Qncorhynchus tshawytscha, were found to be 0.9 and 0.7 mg/L 

nitrite nitrogen in fresh water (Westin, 1974). Smith and 

Williams (1974) tested the effects of nitrite nitrogen and 

observed that yearling rainbow trout, Salmo gairdneri, suffered a 

55 percent mortality after 24 hours at 0.55 mg/L; fingerling 

rainbow trout suffered a 50 percent mortality after 24 hours of 

exposure at 1.6 mg/L; and chinook salmon, Oncorhynchus 

tshawytscha, suffered a 40 percent mortality within 24 hours at 
0.5 mg/L. There were no mortalities among rainbow trout exposed 

to 0.15 mg/L nitrite nitrogen for 48 hours. These data indicate 

that salmonids are more sensitive to nitrite toxicity than are 

other fish species, e.g., minnows, -_-- Phoxinus laevis that 

suffered a 50 percent mortality within 1.5 hours of exposure to 

2,030 mg/L nitrite nitrogen, but required 14 days of exposure 

€or mortality to occur at 10 mg/L (Klingler, 1957), and carp, 

Cyprinus -- care, when raised in a water reuse system, tolerated 

up to 1.8 mg/L nitrite nitrogen (Saeki, 1965). 

Gilbette, et al. (1952) observed that the critical range for 

creek chub, Semotilus - atromaculatus, was 80 to 400 mg/L nitrite 

nitrogen. Wallen et al. (1957) reported a 24-hour LC50 of 1.6 

mg/L nitrite nitrogen, and 48- and 96-hour LC50 values of 1.5 

mg/L nitrite nitrogen for mosquitofish, Gambusia affinis. McCoy 

(1972) tested the nitrite susceptibility of 13 fish species and 

found that logperch, Percina -- caprodes -_I were the most sensitive 

species tested (mortality at 5 mg/L nitrite nitrogen in less 

than 3 hours of exposure) whereas carp, 9prinus carpio, and 

black bullheads, ---- Ictalurus ---- melas 8 survived 40 mg/L nitrite 

nitrogen €or a 48-hour exposure period; the common white sucker, 

Catostomus --..-------I commersoni and the quil lback, Carpiodes cyprinus, 

survived 100 mg/L for 48 and 36 hours, respectively. 

e 

Russo et al. (1974) performed f low-through nitrite bioassays 

in hard water (hardness = 199 mg/L CaC03; alkalinity = 176 mg/L 

CaCQ3; pH = 7.9) on rainbow trout, Salmo gairdneri, of four 

different sizes, and obtained 96-hOUr LC50 values ranging from 

0.19 to 0.39 mg/L nitrite nitrogen. Duplicate bioassays on 12- 

gram rainbow trout were continued long enough for their toxicity 
curves to level off, and asymptotic LC50 concentrations of 

0.14 and 0.15 mg/L were reached in 8 days; on day 19, additional 

mortalities occurred. For 2-gram rainbow trout, the minimum 

tested level of nitrite nitrogen at which no mortalities were 

observed after 10 days was 0.14 mg/L; for the 235-gram trout, the 

minimum level with no mortality after 10 days was 0.06 mg/L. 

It is concluded that (1) levels of nitrate nitrogen at or 

below 90 mg/L would have no adverse effects on warmwater fish 

(Knepp and Arkin, 1973); (2) nitrite nitrogen at or below 5 mg/L 

should be protective of most warmwater fish (McCoy, 1972); and 

(3) nitrite nitrogen at or below 0.06 mg/L should be protective 

of salmonid fishes (Russo et al. 1974; Russo and Thurston, 

1975). These levels either are not known to occur or would be 

unlikely to occur in natural surface waters. 

I 

Recognizing that concentrations of nitrate or nitrite that 

would exhibit toxic effects on warm- or coldwater fish could 

rarely occur in nature, restrictive criteria are not recommended. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
NITROBENZENE 

CRITERIA: 

Aquatic Life 

The available data for nitrobenzene indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 27,000 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No definitive 

data are available concerning the chronic toxicity of 

nitrobenzene to sensitive freshwater aquatic life. 

The available data for nitrobenzene indicate that 

acute toxicity to saltwater aquatic life occurs at 

concentrations as low as 6,680 ug/L and would occur at lower 

concentrations among species thaf; are more sensitive than those 

tested. No definitive data are available concerning the chronic 

toxicity of nitrobenzene to sensitive saltwater aquatic life. 

Human Health 

For comparison purposes, two approaches were used to derive 

criterion levels for nitrobenzene. Based on available toxicity 

data, to protect public health the derived level is 19.8 mg/L. 

using available organoleptic data, to control undesirable taste 

and odor qualities of ambient water the estimated level is 30 

ug/L. It should be recognized that organoleptic data have 

limitations as a basis for establishing a water quality 

criterion, and have no demonstrated relationship to potential 

adverse human health effects. 

(45 F.R. 79318, November 28, 1980) 

NOTE: The U.S. EPA is currently developing Acceptable Daily 

Intake (ADI) or Verified Reference Dose (RID) values for 

Agency-wide use for this chemical. The new value should 

SEE APPENDIX B FOR METHODOLOGY 
be substituted when it becomes available. The January, 

1986, draft Verified Reference Dose document cites an RfD 

of .0005 mg/kg/day for nitrobenzene. 
NITROPHENOLS 

CRITERIA: 

Aquatic Life 

The available data for nitrophenols indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 230 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of nitrophenols to 

sensitive freshwater aquatic life but toxicity to one species of 

algae occurs at concentrations as low as 150 ug/L. 

The available data for nitrophenols indicate that acute 

toxicity to saltwater aquatic life occurs at concentrations as 

low as 4,850 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of nitrophenols to 

sensitive saltwater aquatic life. 

1 

Human Health 

Because of insufficient available data for mono- and 

trinitrophenols, satisfactory criteria cannot be derived at this 

time, using the present guidelines. 

For the protection of human health from the toxic properties 

of dinitrophenols and 2,4-dinitro-o-cresol ingested through 

water and contaminated aquatic organisms, the ambient water 

criteria are determined to be 70 ug/L and 13.4 ug/L, respectively. 

For the protection of human health from the toxic properties 

of dinitrophenols and 2,4-dinitro-o-cresol ingested through 

contaminated aquatic organisms alone, the ambient water criteria 

- 
are determined to be 14.3 mg/L and 765 ug/L, respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
NITROSAMINES 

Aquatic Life 

The available data for nitrosamines indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 5,850 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of nitrosamines to 

sensitive freshwater aquatic life. 

The available data for nitrosamines indicate that acute 

toxicity to saltwater aquatic life occurs at concentrations as 

low as 3,300,000 ug/L and would occur at lower concentrations 

among species that are more sensitive than those tested. No data 

are available concerning the chronic toxicity of nitrosamines to 

sensitive saltwater aquatic life. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to N-nitrosodiethylamine and all 

other nitrosamines except those listed below, through ingestion 

of contaminated water and contaminated aquatic organisms, the 

ambient water concentrations should be zero, based on the non 

threshold assumption for this chemical. However, zero level may 

not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

lifetime are estimated at lo"*, and 1OpcPs The 

corresponding recommended criteria are 8.0 ng/L, 0.8 ng/L, and 

0.08 ng/L, respectively. If these estimates are made for 

0 

L 
consumption of aquatic organisms only, excluding consumption of 

water, the levels are 12,400 ng/L, 1,240 ng/L, and 124 ng/L, 

respectively. 

For the maximum protection of human health from the potential 

carcinoqenic effects of exposure to N-nitrosodimethylamine 

through ingestion of contaminated water and contaminated 

aquatic organisms, the ambient water concentrations should be 

zero, based on the nonthreshold assumption for this chemical. 

However, zero level may not be attainable at the present time. 

Therefore, the levels which may result in incremental increase of 

cancer risk over the lifetime are estimated at 10- 

6, and The corresponding recommended criteria are 14 

ng/L, 1.4 ng/L, and 0.14 ng/L, respectively. If these estimates 

are made for consumption of aquatic organisms only, excluding 

consumption of water, the levels are 160,000 ng/L, 16,000 ng/L, 

and 1,600 ng/L, respectively. 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to N-nitrosodibutylamine through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, 

based on the non threshold assumption for this chemical. 

However, zero level may not be attainable at the present time. 

Therefore, the levels which may result in incremental increase of 

cancer risk over the lifetime are estimated at 10- 

6, and The corresponding recommended criteria are 64 

ng/L, 6.4 ng/L, and 0.64 ng/L, respectively. If these 

estimates are made for consumption of aquatic organisms only, 

excluding consumption of water, the levels are 5,868 ng/L, 587 
ng/L, and 58.7 ng/L, respectively. 0 For the maximum protection o'f human health from the potential 

carcinogenic effects of exposure to N-nitrosopyrrolidine through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

. over the lifetime are estimated at and 

The corresponding recommended criteria are 160 ng/L, 16 ng/L, and 

1.6 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 919,000 ng/L, 91,900 ng/L, and 9,190 ng/L, 

respectively. 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to N-nitrosodiphenylamine 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, 

based on the non threshold assumption for this chemical. 

However, zero level may not be attainable at the present time. 

Therefore, the levels which may result in incremental increase 

of cancer risk over the lifetime are estimated at 10-5, 

The corresponding recommended criteria are 

49,000 ng/L, 4,900 ng/L, and 490 ngfL, respectively. If these 

estimates are made for consumption of aquatic organisms only, 

excluding consumption of water, the levels are 161,000 

ng/L, 16,100 ng/L, and 1,610 ng/L, respectively. 

and 

L, 
(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
-

OIL - AND GREASE 

For domestic water supply: Virtually free from 

oil and grease, particularly from the tastes 

and odors that emanate from petroleum products. 

For aquatic life: 

(1) 0.01 of the lowest continuous flow 96-hour 

LC50 to several important freshwater and 

marine species, each having a demonstrated 

high susceptibility to oils and 

petrochemicals. 

(2) Levels of oils or petrochemicals in the 

sediment which cause deleterious effects to 

the biota should not be allowed. 

(3) Surface waters shall be virtually free from 

floating nonpetroleum oils of vegetable or 

animal origin, as well as petroleum-derived 

oils. 1 

INTRODUCTION: 

It has been estimated that between 5 and 10 million metric 

tons of oil enter the marine environment annually (Blumer, 1970). 

A major difficulty encountered in the setting of criteria for 

oil and grease is that these are not definitive chemical 

categories, but include thousands of organic compounds with 

varying physical, chemical, and toxicological properties. They 

may be volatile or nonvolatile, soluble or insoluble, persistent 

or easily degraded. 

RATIONALE : 

Field and laboratory evidence have demonstrated both acute 

lethal toxicity and long-term sublethal toxicity of oils to 

aquatic organisms. Events such as the Tampico Maru wreck of 

1957 in Baja, California, (Diaz-Piferrer, 1962), and the No. 2 

fuel oil spill in West Falmouth, Massachusetts, in 1969 ..I 
(Hampson and Sanders, 1969), both of which caused immediate death 

to a wide variety of organisms, are illustrative of the lethal 

toxicity that may be attributed to oil pollution. Similarly, a 

gasoline spill in South Dakota in November 1969 (Bugbee and 

Walter, 1973) was reported to have caused immediate death to the 

majority of freshwater invertebrates and 2,500 fish, 30 percent 

of which were native species of trout. Because of the wide 

range of compounds included in the category of oil, it is 

impossible to establish meaningful 96-hour LC50 values for oil 

and grease without specifying the product involved. 

However, as the data in Table 6 show, the most susceptible 

category of organisms, the marine larvae, appear to be intolerant 

of petroleum pollutants, particularly the water soluble 

compounds, at concentrations as low as 0.1 mg/L. 

The long-term sublethal effects of oil pollution refer to 

interferences with cellular and physiological processes such as 

feeding and reproduction and do not lead to immediate death of 

the organism. Disruption of such behavior apparently can result 

from petroleum product concentrations as low as 10 to 100 ug/L 

(see Table 7). 

Table 7 summarizes some of the sublethal toxicities for 

various petroleum pollutants and aquatic species. In addition to 

sublethal effects reported at the 10 to 100 ug/L level, it has 

been shown that petroleum products can harm aquatic life at 

concentrations as low as 1 ug/L (Jacobson and Boylan, 1973). 

Bioaccumulation of petroleum products presents two especially 

important public health problems: (1) the tainting of edible, 
aquatic species, and (2) the possibility of edible marine 

organisms incorporating the high boiling, carcinogenic polycyclic 

aromatics in their tissues. Nelson-Smith (1971) reported that 

0.01 mg/L of crude oil caused tainting in oysters. Moore et al. 

(1973) reported that concentrations as low as 1 to 10 ug/L could 

lead to tainting within very short periods of time. It has been 

shown that chemicals responsible for cancer in animals and man 

(such as 3,4-benzopyrene) occur in crude oil (Blumer, 1970). It 

also has been shown that marine organisms are capable of 

incorporating potentially carcinogenic compounds into their body 

fat where the compounds remain unchanged (Blumer, 1970). 

Oil pollutants may also be incorporated into sediments. 

There is evidence that once this occurs in the sediments below 

the aerobic surface layer, petroleum oil can remain unchanged and 

toxic for long periods, since its rate of bacterial degradation 

is slow. For example, Blumer (1970) reported that No. 2 fuel 

oil incorporated into the sediments after the West Falmouth spill 

persisted for over a year, and even began spreading in the form 

of oil-laden sediments to more distant areas that had remained 

unpolluted immediately after the spill. The persistence of 

unweathered oil within the sediment could have a long-term effect 

on the structure of the benthic community or cause the demise of 

specific sensitive important species. Moore et al. (1973) 

reported concentrations of 5 mg/L for the carcinogen 3, 4- 

benzopyrene in marine sediments. 

Mironov (1967) reported that 0-01 mg/L oil produced deformed 

and inactive flatfish larvae. Mironov (1970) also reported 

inhibition or delay of cellular division in algae by oil 

-, 
concentrations of to 10-1 mg/L. Jacobson and Boylan (1973) 

reported a reduction in the chemotactic perception of food by the 

snail, Nassarius -- obsoletus at kerosene concentrations of 0.001 

to 0.004 mg/L. Bellen et al. (1972) reported decreased survival 

and fecundity in worms at concentrations of 0.01 to 10 mg/L of 

detergent. 

Because of the great variability in the toxic properties of 

oil, it is difficult to establish a numerical criterion which 

would be applicable to all t-jpes of oil. Thus, an application 

factor of 0.01 of the 96-hour LC50 as determined by using 

continuous flow with a sensitive resident species should be 

employed for individual petrochemical components. 

There is a paucity of toxicological data on the ingestion of 

the components of refinery wastewaters by humans or by test 

animals. It is apparent that any tolerable health concentrations 

for petroleum-derived substances far exceed the limits of taste 

and odor. Since petroleum derivatives become organoleptically 

objectionable at concentrations far below the human chronic 

toxicity, it appears that hazards to humans will not arise from 

drinking oil-polluted waters (Johns Hopkins Univ., 1956; Mckee 

and Wolf, 1963). Oils of animal or vegetable origin generally 

are nontoxic to humans and aquatic life. 

In view of the problem of petroleum oil incorporation in 

sediments, its persistence and chronic toxic potential, and the 

present lack of sufficient toxicity data to support specific 

criteria, concentrations of oils in sediments should not approach 

levels that cause deleterious effects to important species or the 
bottom community as a whole. 

Petroleum and nonpetroleum oils share some similar physical 

and chemical properties. Because they share common properties, 

they may cause similar harmful effects in the aquatic 

environment by forming a sheen, film, or discoloration on the 

surface of the water. Like petroleum oils, nonpetroleum oils 

may occur at four levels of the aquatic environment: (a) floating 

on the surface, (b) emulsified in the water column, (c) 

solubilized, and (d) settled on the bottom as a sludge. Analogous 

to the grease balls from vegetable oil and animal fats are the 

tar balls of petroleum origin which have been found in the marine 

environment or washed ashore on beaches. 

Oils of any kind can cause (a) drowning of waterfowl because 

of loss of buoyancy, exposure because of loss of insulating 0 capacity of feathers, and starvation and vulnerability to 

predators because of lack of mobility; (b) lethal effects on fish 

by coating epithelial surfaces of gills, thus preventing 

respiration; (c) potential fishkills resulting from biochemical 

oxygen demand: (d) asphyxiation of benthic life forms when 

floating masses become engaged with surface debris and settle on 

the bottom: and (e) adverse aesthetic effects of fouled 

shorelines and beaches. These and other effects have been 

documented in the U.S. Department of Health, Education and 

Welfare report on Oil Spills Affecting the Minnesota and 

Mississippi Rivers and the 1975 Proceedings of the Joint 

Conference on Prevention and Control of Oil Spills. e Oils of animal or vegetable origin generally are chemically _, 

nontoxic to humans or aquatic life; however, floating sheens of 
such oils result in deleterious environmental effects described 

in this criterion. Thus, it is recommended that surface waters 

shall be virtually free from floating nonpetroleum oils of 

vegetable or animal origin. This same recommendation applies to 

floating oils of petroleum origin since they too may produce 

similar effects. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
DISSOLVED OXYGEN 0 NATIONAL CRITERIA: 

The national criteria for ambient dissolved oxygen concentrations for the protection of freshwater aquatic life are presented 

in Table 1. The criteria are derived from the production impairment estimates which are based primarily upon growth data and 

information on temperature, disease, and pollutant stresses. The 

average dissolved oxygen concentrations selected are values 0.5 

mg/L above the slight production impairment values and represent values between no production impairment and slight 

production impairment. Each criterion may thus be viewed as an 

estimate of the threshold concentration below which detrimental 

effects are expected. 

Criteria for coldwater fish are intended to apply to waters 

containing a population of one or more species in the family 

Salmonidae {Bailey et al., 1970) or to waters containing other 

coldwater or coolwater fish deemed by the user to be closer to 

salmonids in sensitivity than to most warmwater species. 

Although the acute lethal limit €or salmonids is at or below 3 

mg/L, the coldwater minimum has been established at 4 mg/L 

because a significant proportion of the insect species common 

to salmonid habitats are less tolerant of acute exposures to low 

dissolved oxygen than are salmonids. Some coolwater species may 

require more protection than that afforded by the other life 

stage criteria for warmwater fish and it may be desirable to 

protect sensitive coolwater species with the coldwater 

criteria. Many states have more stringent dissolved oxygen 

standards for cooler waters, waters that contain either 

a ._ 
salmonids, nonsalmonid coolwater fish, or the sensitive centrachid, the smallmouth bas5 The warmwater criteria are necessary 

to protect early life stages of warmwater fish as sensitive as 

as channel catfish and to protect other life stages of fish as 

sensitive as largemouth bass. Criteria for early life stages are 

intended to apply only where and when these stages occur. These 

criteria represent dissolved oxygen concentrations which EPA 

believes provide a reasonable and adequate degree of protection 

for freshwater aquatic life. 

The criteria do not represent assured no-effect levels. 

However, because the criteria represent worst case conditions 

(i.e. for wasteload allocation and waste treatment plant design) ~ 

conditions will be better than the criteria nearly all of the 

time at most sites. In situations where criteria conditions are 

just maintained for considerable periods the proposed criteria 

represent some risk of production impairment. This impairment 

would depend on innumerable other factors. If slight production 

impairment or a small but undefinable risk of moderate impairment 

is unacceptable, than one should use the "no production impairment" values given in the document as means and the "slight 

production impairment') values as minima. The table which presents these concentrations is reproduced here as table 2. 

The criteria do represent dissolved oxygen concentrations 

believed to protect the more sensitive populations of organisms 

against potentially damaging production impairment. The 

dissolved oxygen concentrations in the criteria are intended to 

be protective at typically high seasonal environmental temperatures for the appropriate taxonomic and life stage classi- 
0 Table 1. Water quality criteria for ambient dissolved oxygen 

concentration. 

Coldwater Criteria Warmwater Criteria 

Stages Ue Stages Stages Stages 

Other Life Early kife Other Life 

30 Day Mean NA3 6.5 NA 5.5 

7 Day Mean 9.5 (6.5) NA 6.0 NA 

7 Day Mean NA 

Minimum 

5.0 NA 4.0 

1 Day 8.0 (5.0) 4.0 5.0 3.0 

Minimum 415 

These are water column concentrations recommended to achieve 

the required intergravel dissolved oxygen concentrations 

shown in parentheses. The 3 mg/L differential is discussed in 

the criteria document. For species that have early life 

stages exposed directly to the water column, the figures in 

parentheses apply. 

Includes all embryonic and larval stages and all juvenile 

forms to 30-days following hatching. 

0 

3 NA (not applicable). 

For highly manipulatable discharges, further restrictions 

apply (see page 37) 

All minima should be considered as instantaneous 

concentrations to be achieved at all times. 

fications, temperatures which are often higher than those used in 

the research from which the criteria were generated, especially 

for other than early life stages. 

Where natural conditions alone create dissolved oxygen 

concentrations less than 110 percent of the applicable criteria 

means or minima or both, the minimum acceptable concentration is 
90 percent of the natural concentration. These values are 

similar to those presented graphically by Doudoroff and 

Shumway (1970) and those calculated from Water Quality Criteria 

1972 (NAS/NAE, 1973). Absolutely no anthropogenic dissolved 

oxygen depression in the potentially lethal area below the 

1-day minima should be allowed unless special care is taken to 

ascertain the tolerance of resident species to low dissolved 

oxygen. 

.If daily cycles of dissolved oxygen are essentially 

sinusoidal, a reasonable daily average is calculated from the 

day's high and low dissolved oxygen values. A time-weighted 

average may be required if the dissolved oxygen cycles are 

decidedly non-sinusoidal. Determining the magnitude of daily 

dissolved oxygen cycles requires at least two appropriately timed measurements daily, and characterizing the shape of the 

cycle requires several more appropriately spaced measurements. 

Once a series of daily mean dissolved oxygen concentrations 

are calculated, an average of these daily means can be calculated (Table 3). For embryonic, larval, and early life stages, 

the averaging period should not exceed 7 days. This short time 

is needed to adequately protect these often short duration, most 

sensitive life stages. Other life stages can probably be 

adequately protected by 30-day averages. Regardless of the 

averaging period, the average should be considered a moving 

average rather than a calendar-week or calendar-month average. 
0 Table 2. Dissolved Oxygen Concentrations (mg/L) Versus 

Quantitative Level of Effect. 

1. Salmonid Waters 

a. Embryo and Larval Stages 

No Production Impairment = 11* (8) 

Slight Production Impairment = 9* (6) 

Moderate Production Impairment = 8* (5) 

Severe Production Impairment = 7* (4) 

Limit to Avoid Acute Mortality = 6* (3) 

(* Note: These are water column concentrations recommended to 

achieve the required intergravel dissolved oxygen 

concentrations shown in parentheses. The 3 mg/L 

difference is discussed in the criteria document.) 

b. Other Life Stages 

No Production Impairment =8 

light Production Impairment =6 

Moderate Production Impairment = 5 

Severe Production Impairment = 4 

Limit to Avoid Acute Mortality = 3 

2. Nonsalmonid Waters 

a. Early Life Stages 

No Production Impairment = 6.5 

Slight Production Impairment = 5.5 

Moderate Production Impairment = 5 

Severe Production Impairment = 4,5 

Limit to Avoid Acute Mortality = 4 

b. Other Life Stages 

No Production Impairment =6 

Slight Production Impairment = 5 

Moderate Production Impairment = 4 

Severe Production Impairment = 3.5 

Limit to Avoid Acute Mortality = 3 

3. Invertebrates 

No Production Impairment =8 

Some Production Impairment =5 

Acute Mortality Limit =4 
Table 3. Sample calculations for determining daily means 

and 7-day mean dissolved oxygen concentrations (30-day 

averages are calculated in a similar fashion using 30 

days data). 

Dissolved Oxygen (mg/L) 

Day Daily Max. Daily Min. Daily Mean 

9.0 7.0 

10.0 7.0 

11.0 8.0 

12. oa 8.0 

10.0 8.0 

11.0 90 

12.03 10.0 

8.0 

8.5 

9.5 

9.0 

10.0 

1Q.5c 

9.5b 

57.0 65.0 

1-day Minimum 7.0 

?-day Mean Minimum 8.1 

?-day Mean 9.3 

a Above air saturation concentration (assumed to be 11.0 

mg/L for this example). 

b (11.0 -b 8.0)2. 

c (11 0 +10.0)2. 

The criteria have been established on the basis that the 

maximum dissolved oxygen value actually used in calculating any 

daily mean should not exceed the air saturation value. This 

consideration is based primarily on analysis of studies of 

cycling dissolved oxygen and the growth of largemouth bass 

(Stewart et al., 1967), which indicated that high dissolved 

oxygen levels (> 6 mg/L) had no beneficial effect on growth. 

During periodic cycles of dissolved oxygen concentrations, 

minima lower than acceptable constant exposure levels are tolerable so long as: 
1. the average concentration attained meets or exceeds the 

2. the average dissolved oxygen concentration is calculated as 

3. 

criterion; 

recommended in Table 3; and 

the minima are not unduly stressful and clearly are not lethal. 

A daily minimum has been included to make certain that no 

acute mortality of sensitive species occurs as a result of lack 

of oxygen. Because repeated exposure to dissolved oxygen 

concentrations at or near the acute lethal threshold will be 

stressful and because stress can indirectly produce mortality or 

other adverse effects (e-g., through disease), the criteria 

are designed to prevent significant episodes of continuous or 

regularly recurring exposures to dissolved oxygen concentrations 

at or near the lethal threshold. This protection has been 

achieved by setting the daily minimum for early life stages at 

the subacute lethality threshold, by the use of a 7-day averaging 

period for early life stages, by stipulating a 7-day mean minimum 

value for other life stages, and by recommending additional 

limits for manipulatable discharges. 

The previous EPA criterion for dissolved oxygen published in 

Quality Criteria for Water (USEPA, 1976) was a minimum of 5 mg/L 

(usually applied as a 7410) which is similar to the current 

criterion minimum except for other life stages of warmwater fish 

which now allows a 

criteria are similar 

of the Federal Water 

7-day mean minimum of 4 mg/L. The new 

to those contained in the 1968 "Green Book" 

Pollution Control Federation (FWPCA, 1968). 
A. The Criteria and Monitoring and Design Conditions 

The acceptable mean concentrations should be attained most of 

the time, but some deviation below these values would probably 

not cause significant harm. Deviations below the mean will 

probably be serially correlated and hence apt to occur on 

consecutive days. The significance of deviations below the mean 

will depend on whether they occur continuously or in daily 

cycles, the former being more adverse than the latter. Current 

knowledge regarding such deviations is limited primarily to laboratory growth experiments and by extrapolation to other activityrelated phenomena. 

Under conditions where large daily cycles of dissolved oxygen 

occur, it is possible to meet the criteria mean values and 

consistently violate the mean minimum criteria. Under these 

conditions the mean minimum criteria will clearly be the 

limiting regulation unless alternatives such as nutrient 

control can dampen the daily cycles. 

The significance of conditions which fail to meet the 

recommended dissolved oxygen criteria depend largely upon five 

factors: (1) the duration of the event; (2) the magnitude of the 

dissolved oxygen depression; (3) the frequency of recurrence; (4) 

the proportional area of the site failing to meet the criteria, 

and (5) the biological significance of the site where the event 

occurs. Evaluation of an event's significance must be largely 

case- and site-specific. Common sense would dictate that the 

magnitude of the depression would be the single most important 

factor in general, especially if the acute value is violated. A 
logical extension of these considerations is that the event must 

be considered in the context of the level of resolution of the 

monitoring or modeling effort. Evaluating the extent, duration, 

and magnitude of an event must be a function of the spatial and 

temporal frequency of the data. Thus, a single deviation below 

the criterion takes on considerably less significance where 

continuous monitoring occurs than where sampling is comprised of once-a-week grab samples. This is so because based on 

continuous monitoring the event is provably small, but with 

the much less frequent sampling the event is not provably small 

and can be considerably worse than indicated by the sample. The 

frequency of recurrence is of considerable interest to those 

modeling dissolved oxygen concentrations because the return 

period, or period between recurrences, is a primary modeling 

consideration contingent upon probabilities of receiving water 

volumes, waste loads, temperatures, etc. It should be apparent 

that return period cannot be isolated from the other four factors 

discussed above. Ultimately, the question of return period may 

be decided on a site-specific basis taking into account the 

other factors (duration, magnitude, areal extent, and biological significance) mentioned above. Future studies of temporal 

patterns of dissolved oxygen concentrations, both within and 

between years, must be conducted to provide a better basis for 

selection of the appropriate return period. 

In conducting wasteload a1 location and treatment plant design 

computations, the choice of temperature in the models will be 

important. Probably the best option would be to use temperatures 

consistent with those expected in the receiving water over the 
critical dissolved oxygen period for the biota. 

B. The Criteria and Manipulatable Discharges 

If daily minimum DOs are perfectly serially correlated, 

i.e, if the annual lowest daily minimum dissolved oxygen concentration is adjacent in time to the next lower daily minimum 

dissolved oxygen concentration and one of these two minima is 

adjacent to the third lowest daily minimum dissolved oxygen 

concentration, etc., then in order to meet the 7-day mean 

minimum criterion it is unlikely that there will be more than 

three or four consecutive daily minimum values below the acceptable 7-day mean minimum. Unless the dissolved oxygen pattern is 

extremely erratic, it is also unlikely that the lowest 

dissolved oxygen concentration will be appreciably below 

the acceptable 7-day mean minimum or that daily minimum values 

below the 7-day mean minimum will occur in more than one or two 

weeks each year. For some discharges, the distribution of 

dissolved oxygen concentrations can be manipulated to varying 

degrees. Applying the daily minimum to manipulatable discharges 

would allow repeated weekly cycles of minimum acutely acceptable 

dissolved oxygen values, a condition of unacceptable stress 

and possible adverse biological effect. For this reason, the 

application of the one day minimum criterion to manipulatable 

discharges must limit either the frequency of occurrence of 

values below the acceptable 7-day mean minimum or must impose 

further limits on the extent of excursions below the 7-day mean 

minimum. For such controlled discharges, it is recommended that 

the occurrence of daily minima below the acceptable 7-day mean 

I 
minimum be limited to 3 weeks per year or that the acceptable 

one-day minimum be increased to 4.5 mg/L for coldwater fish and 

3.5 mg/L for warmwater fish, Such decisions could be sitespecific based upon the extent of control and serial correlation. 

e 
PARATHION 

CRITERION: 

0.04 ug/L for freshwater and marine aquatic life. 

RATIONALE: 

Acute static LC50 values of the organophosphorus pesticide, 

parathion, for freshwater fish have ranged generally from about 

50 ug/L for more sensiitive species such as bluegills, Lepomis 

macrochi=, to about 2.5 mg/L for the more resistant species 

such as minnows (U.S. Environ. Prot. Agency, 1975). In flowing 

water exposures, Spacie (1975) obtained 96-hour LC50 values of 

0.5 mg/L, 1.6 mg/L, and 1.76 mg/L f o r b l u e g i l l s , _- Lepomis ____ 

macrochirus, fathead minnows, -- Pimep&ales promelas, and brook 

trout, Salvelinus _______I fontinalis respectively. Korn and Earnest 

(1974) found a 96-hOUr LC50 of 18 ug/L for juvenile freshwater 

and estuarine striped bass, Morone saxatilis, in a flowing water 

system, 

I 

Few chronic exposure data are available for aquatic 

organisms. Brown bullheads, fctalurus nebulosus, exposed to 30 

ug/L parathion for 30 days exhibited tremors; at 60 ug/L they 

convulsed and were found to have developed a deformed vertebral 

column (Mount and Boyle, 1969). In a 23-mOnth exposure of 

bluegills, Spacie (1975) observed deformities (scoliosis and a 

characteristic protrusion in the throat region) at 0.34 ug/L, but 

not at 0.16 ug/L. Tremors, convulsions, hypersensitivity, and 

hemorrhages also were evident at higher concentrations. 

Reproductive impairment and deformities were observed in 

fathead minnows exposed to 4.0 ug/L for 8 1/2 months. 
Development of brook trout, S, fontinalis -- embryos exposed to 32 

ug/L was abnormal and mortalities associated with premature 

hatching were observed. Embryos at 10 ug/L appeared normal. No 

adverse effects on juveniles and adults was evident during 9 

months' exposure to 7 ug/L. 

Inhibition of cholinesterase enzymes is the well-established 

mode of physiological action of parathion and other organic 

phosphorus pesticides (Weiss, 1958). The degree of inhibition of 

brain acetylcholinesterase (AChE) activity has been the most 

frequently used measure of effect of these pesticides. Various 

studies (Weiss, 1958, 1959, 1961; Murphy et al., 1968; Gibson et 

al. 1969) have shown the degree of inhibition to be dependent 

upon toxicant concentration, length of exposure, and species 

sensitivity. The results of these studies have also indicated 

AChE inhibition ranging from 25 to 90 

(1959) also showed that susceptibility 

of recovery of AChE activity following 

that death results from 

percent of normal. Weiss 

depended upon the extent 

prior exposure and that the recovery period for fish exposed to 

parathion was relatively long. In bluegills, AChE activity was 

only 50 percent recovered 30 days after exposure to 1 mg/L for 6 

to 7 hours (Weiss, 1961). 

Some of the other physiological effects observed to result 

from exposure of fish to parathion have been inhibition of 

spermatogenesis in guppies, _______- Poecilia reticulata at 10 ug/L 

(Billard and deKinkelin, 1970), alternation of oxygen 

consumption rate in bluegills, Lepomis macrochirus, at 100 ug/L 

(Dowden, 1966), and liver enlargement associated with increased 

pesticide-hydrolizing capability in mosquitof ish, Gambusia 
0 affinis (Ludke, 1970). 

Parathion has been found acutely toxic to aquatic 

invertebrates at under 1 ug/L e.g., a 50-hour LC50 of 0.8 ug/L 

for - Daphnia maqgg; 48-hour LC50 of 0.6 ug/L for -- Daphnia -- pulex; --- 

48-hour LC50 of 0.37 for Simocephalus ---_ _____-_ serrulatus (a daphnid) 

(Sanders and Cope, 1966); a 5-day LC50 of 0.93 ug/L for the 

larval stonefly, Acroneuria pacifica (Jensen and Gaufin, 1964); 

and a 96-hour LC50 of 0.43 ug/L for the larval caddisfly 

- HydroEsychc --- - -- californica - -------- (Gaufin et al. 1965). Mulla and 

Khasawinah (1969) obtained a 24-hour LC50 of 0.5 ug/L for 4th 

instar larvae of the midge -- Tanypus - qrodhausi. Spacie (1975) 

obtained 96-hour LC5O's in flow-through bioassays of 0.62 ug/L 

for Daphnia magna, 0.40 ug/L for the scud, Gammarus fasciatus 

and 31.0 ug/L for 4th instar of Chironomous tentans, a midge. 

Other invertebrates have been found acutely sensitive to 

parathion in concentrations of from 1 to 30 ug/L in water (U.S. 

Environ. Prot. Agency, 1975). 

Few longer exposures have been conducted. Jensen and Gaufin 

(1964) obtained 30-day LC50's for Pteronarcys - ----I-- californica and 

----_ Acroneuria pacifica ---- of 2.2 and 0.44 ug/L, respectively. Spacie 

(1975) found the 3-week LC50 for Daphnia mxna - to be 0.14 ug/L. 

Statistically significant reproductive impairment occurred at 

concentrations above 0.08 ug/L. A 43-day LC50 of 0.07 ug/L was 

reported for Gammarus fasciatus and a concentration of 0.04 ug/L 

produced significantly greater mortality than among controls. 

Limited information is available on persistence of parathion 

~ in water. Eichelberger and Lichtenberg (1971) determined the 
half-life in river water (pH 7.3 - 8.0) to be 1 week. Using AChE 

inhibitory capacity as the indicator, Weiss and Gakstatter 

(1964) found the half-life of parathion or its active breakdown 

products to be 40, 35, and 20 days in ‘lnaturalll waters having a 

pH of 5.1. 7.0, and 8.4, respectively. The possibility of 

breakdown resulting in compounds more toxic than parathion was 

suggested by Burke and Ferguson (1969) who determined that the 

toxicity of this pesticide to mosquitofish 1 Gambusia ---_I affinis 

was greater in static than in flowing water test systems. 

Sanders (1972), in 96-hour bioassays with the scud, Gammarus _- 

-------I fasciatus and glass shrimp, Palaemonetes -_-I_kadiakensis I also 

observed greater toxicity under static than in flow-through 

conditions. 

a Tissue accumulations of parathion by exposed aquatic 

organisms are not great and do not appear to be very persistent. 

Mount and Boyle (1969) observed concentrations in the blood of 

bullhead, --------- Ictalurus -----I melas up to about 50 times water 

concentrations. Spacie (1975) found muscle concentrations in 

chronically exposed brook trout, - S. ___--_I fontinalis to be several 

hundred times water concentrations; bluegills, -- Lepomis ---- 

macrochirus, had about 25 times water concentrations in their 

bodies. Leland (1968) demonstrated a biological half-life of 

parathion in rainbow trout, Salmo Eirdneri 1 exposed and then 

placed in fresh water to be only 30 to 40 hours. It is not 

expected that parathion residues in aquatic organisms exposed to 

the recommended criterion concentrations will be a hazard to 

consumer organisms. a 
Weiss and Gakstatter (1964) have shown that 15-day continuous 

exposure to parathion (1.0 ug/L) can produce progressively 

greater (i.e-, cumulative) brain AChE inhibition in a fish 

species. After substantial inhibition by parathion exposure, it 

takes several weeks for brain AChE of exposed fishes to return 

to normal even though exposure is discontinued (Weiss, 1959, 

1961). fishes by 46 percent or more 

has been associated with harmful effects in exposures to 

organophosphate pesticides for one life cycle (Eaton, 1970) and 

for short periods (Carter, 1971; Coppage and Duke, 1971; Coppage, 

1972; Coppage and Matthews, 1974; Post and Leasure, 1974; Coppage 

et al. 1975). It has been shown that a concentration of 10 ug 

Inhibition of brain AChE of 

parathion/L of flowing seawater kills 40 to 60 percent of the 

marine fishes &agodon ---- --rhomboides -------- (pinfish) and Leostomus -- 

xanthurus __ (spot) in 24 hours and causes about 87 to 92 percent 

brain AChE inhibition (Coppage and Matthews, 1974.) Similar 

inhibition of AChE and mortality were caused in sheepshead 

minnows, Cyprinodon variegatus, in 2, 24, 48, and 72 hours at 

concentrations of 5,000, 2,000, 100, and 10 ug/L, respectively in 

static tests (Coppage, 1972). These data indicate that 

reductions of brain AChE activity of marine fishes by 70 to 80 

percent or more in short-term exposures to parathion may be 

associated with some deaths. 
0 Other estimates of parathion toxicity to marine organisms 

follow. The 48-hour EC50 for parathion to Penaeus duorarum was 

found to be 0.2 ug/L (Lowe et al. 1970). Lahav and Sarig (1969) 

reported the 96-hour LC50 for mullet, Mugil cephalus to be 125 

ug/L. The shell growth of the oyster, Crassostrea virginica, was 

found by Lowe et al. (1970) to be decreased by 22 percent after 

96 hours in 1.0 mg/L. 

An application factor of 0.1 is applied to the 96-hour LC50 

data for invertebrates which range upward from 0.4 ug/L. A 

criteria of 0.04 ug/L is recommended for marine and freshwater 

aquatic life. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
0 

CRITERIA: 

PENTACHLOROPHENOL 

Aquatic Life 

The available data for pentachlorophenol indicate that acute 

and chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 55 and 3.2 ug/L, respectively, and would 

occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for pentachlorophenol indicate that acute 

and chronic toxicity to saltwater aquatic life occur at 

concentrations as low as 53 and 34 ug/L, respectively, and would 

occur at lower concentrations among species that are more 

sensitive than those tested. 

Human Health 

For comparison purposes, two approaches were used to derive 

criterion levels for pentachlorophenol. Based on available 

toxicity data, to protect public health the derived level is 1.01 

W/L. Using available organoleptic data, to control 

undesirable taste and odor qualities of ambient water the 

estimated level is 30 ug/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a 

water quality criterion, and have no demonstrated 

relationship to potential adverse human health effects. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERIA: 

Range 

5 - 9 Domestic water supplies (welfare) 

6.5 - 9.0 Freshwater aquatic life 

6.5 - 8.5 Marine aquatic life (but not more than 

0.2 units outside of normal lyoccurring 

range. ) 

INTRODUCTION : 

I8pH" is a measure of the hydrogen ion activity in a water 

sample. It is mathematically related to hydrogen ion activity 

according to the expression: pH = -log 10 (H'), where (H') is the 

hydrogen ion activity. 

The pH of natural waters is a measure of acid-base 

equilibrium achieved by the various dissolved compounds, salts, 

and gases. The principal system regulating pH in natural waters 

is the carbonate system which is composed of carbon dioxide 

(C02), carbonic acid, (HzC03), bicarbonate ion (HC03) and 

carbonate ions (Cog). The interactions and kinetics of this 

system have been described by Stumm and Morgan (1970). 

pH is an important factor in the chemical and biological 

systems of natural waters. The degree of dissociation of weak 

acids or bases is affected by changes in pH. This effect is 

important because the toxicity of many compounds is affected by 

the degree of dissociation. One such example is hydrogen cyanide 

(HCN). Cyanide toxicity to fish increases as the pH is lowered - 

because the chemical equlibrium is shifted toward an increased 

concentration of HCN. Similar results have been shown for 

hydrogen sulfide (Has) (Jones, 1964). 
The solubility of metal compounds contained in bottom 

sediments or as suspended material, also is affected by pH. For 

example, laboratory equilibrium studies under anaerobic 

conditions indicated that pH was an important parameter involved 

in releasing manganese from bottom sediments (Delfino and Lee, 

1971). 

The pH of a water does not indicate ability to neutralize 

additions of acids or bases without appreciable change. This 

characteristic, termed "buffering capacity," is control led by the 

amounts of alkalinity and acidity present. 

RATIONALE: 

Knowledge of pH in the raw water used €or public water 

supplies is important because without adjustment to a suitable 

level, such waters may be corrosive and adversely affect 

treatment processes including coagulation and chlorination. 

Coagulation for removal of colloidal color by use of aluminum 

or iron salts generally has an optimum pH range of 5.0 to 6-5 

(Sawyer, 1960). Such optima are predicated upon the availability 

of sufficient alkalinity to complete the chemical reactions. 

The effect of pH on chlorine in water principally is on the 

equilibrium between hypochlorous acid (HOC1) and the hypochlorite 

ion (OC1-) according to the reaction: 

HOCl = H+ + OC1- 

Butterfield (1984) has shown that chlorine disinfection is more 

effective at values less than pH 7. Another study (Reid and 

Carlson, 1974) has indicated, however, that in natural waters no 

c significant difference in the kill rate for Escherichia -- coli was 
observed between pH 6 and pH 8. 

corrosion of plant equipment and piping in the distribution 

system can lead to expensive replacement as well as the 

introduction of metal ions such as copper, lead, zinc, and 

cadmium. Langelier (1936) developed a method to calculate and 

control water corrosive activity that employs calcium carbonate 

saturation theory and predicts whether the water would tend to 

dissolve or deposit calcium carbonate. By maintaining the pH 

at the proper level, the distribution system can be provided with 

a protective calcium carbonate lining which prevents metal pipe 

corrosion. Generally, this level is above pH 7 and frequently 

approaches pH 8.3, the point of maximum bicarbonate/carbonate 

buffering. 

Since pH is relatively easily adjusted prior to and during 

water treatment, a rather wide range is acceptable for waters 

serving as a source of public water supply. A range of p~ from 

5.0 to 9.0 would provide a water treatable by typical 

(coagulation, sedimentation, filtration, and chlorination) 

treatment plant processes. As the range is extended, the cost of 

neutralizing chemicals increases. 

A review of the effects of pH on fresh water fish has been 

published by the European Inland Fisheries Advisory Commission 

(1969). The commission concluded: 

There is no definite pH range within which a fishery is 

unharmed and outside which it is damaged, but rather, there 

is a gradual deterioration as the pH values are further 

removed from the normal range. The pH range which is not 

directly lethal to fish is 5 - 9; however, the toxicity of 

several common pollutants is markedly affected by pH changes 

within this range, and increasing acidity or alkalinity may 

make these poisons more toxic. Also, an acid discharge may 

liberate sufficient C02 from bicarbonate in the water either 
to be directly toxic, or to cause the pH range 5 - 6 to 

become lethal. 

Mount (1973) performed bioassays on the fathead minnow, 

Pimephal- promelas, for a 13-mQnth, one generation time period 

to determine chronic pH effects. Tests were run at pH 

levels of 4.5, 5.2. 

PH 

Range Effect on Fish* 

5.0 - 6.0 Unlikely to be harmful to any species unless either the 

concentration of free C02 is greater than 20 ppm, or 

the water contains iron salts which are precipitated as 

ferric hydroxide, the toxicity of which is not known. 

is present in excess of 100 ppm. 

6.0 - 6.5 Unlikely to be harmful to fish unless free carbon dioxide 

6.5 - 9.0 Harmless to fish, although the toxicity of other poisons 

may be affected by changes within this range. 

EIFAC, 1969 
5.9, 6.6, and a control of 7.5. At the two lowest pH values (4.5 

and 5.2) behavior was abnormal and the fish were deformed. At pH 

values less than 6.6, egg production and hatchability were 

reduced when compared with the control. It was concluded that a 

pH of 6.6 was marginal for vital life functions. 

Bell (1971) performed bioassays with nymphs of caddisflies 

(two species) stonef lies (four species), dragonflies (two 

species), and mayflies (one species). All are important fish 

food organisms. The 30-day TL50 values ranged from 2.45 to 5.38 

with the caddisflies being the most tolerant and the mayflies 

being the least tolerant. The pH values at which 50 percent of 

the organisms emerged ranged from 4.0 to 6.6 with increasing 

percentage emergence occurring with the increasing pH values. 

Based on present evidence, a pH range of 6.5 to 9.0 appears 

to provide adequate protection for the life of freshwater fish 

and bottom dwelling invertebrates fish food organisms. Outside of 

this range, fish suffer adverse physiological effects increasing 

in severity as the degree of deviation increases until lethal 

levels are reached. 

Conversely, rapid increases in pH can cause increased NH3 

concentrations that are also toxic. Ammonia has been shown to be 

10 times as toxic at pH 8.0 as at pH 7.0 (EIFAC, 1969). 

The chemistry of marine waters differs from that of fresh 

water because of the large concentration of salts present. In 

addition to alkalinity based on the carbonate system, there is 

also alkalinity from other weak acid salts such as borate. 

Because of the buffering system present in seawater, the 
naturally occurring variability of pH is less than in fresh 

water. Some marine communities are more sensitive to pH change 

than others (NAS, 1974). Normal pH values in seawater are 8.0 to 

8.2 at the surface, decreasing to 7.7 to 7.8 with increasing 

depth (Capurro, 1970). The NAS Committee's review (NAS, 1974) 

indicated that plankton and benthic invertebrates are probably 

more sensitive than fish to changes in pH and that mature forms 

and larvae of oysters are adversely affected at the extremes of 

the pH range of 6.5 to 9.0. However, in the shallow, 

biologically active waters in tropical or subtropical areas, 

large diurnal pH changes occur naturally because of 

photosynthesis. pH values may range from 9.5 in the daytime to 

7.3 in the early morning before dawn. Apparently, these 

communities are adapted to such variations or intolerant species 

are able to avoid extremes bymoving out ofthearea. 

For open ocean waters where the depth is substantially 

greater than the euphotic zone, the pH should not be changed more 

than 0.2 units outside of the naturally occurring variation or in 

any case outside the range of 6.5 to 8.5. For shallow, highly 

productive coastal and estuarine areas where naturally occurring 

variations approach the lethal limits for some species, changes 

in pH should be avoided, but in any case not exceed the limits 

established for fresh water, i.e., pH of 6.5 to 9.0. As with 

freshwater criteria, rapid pH fluctuations that are caused by 

waste discharges should be avoided. Additional support for these 

limits is provided by Zirino and Yamamoto (1972). These 0 investigators developed a model which illustrates the effects of %. 

variable pH on copper, zinc, cadmium, and lead; small changes in 
p~ cause large shifts in these metallic complexes. Such changes 

may affect toxicity of these metals. 

For the industrial classifications considered, the NAS report 

(NAs, 1974) tabulated the range of pH values used by industry for 

various process and cooling purposes. In general, process waters 

used varied from pH 3.0 to 11.7, while cooling waters used varied 

from 5.0 to 8.9. Desirable pH values are undoubtedly closer to 

neutral to avoid corrosion and other deleterious chemical 

reactions. Waters with pH values outside these ranges are 

considered unusable for industrial purposes. 

The pH of water applied for irrigation purposes is not 

normally a critical parameter. Compared with the large buffering 

capacity of the soil matrix, the pH of applied water is rapidly 

changed to approximately that of the soil. The greatest danger 

in acid soils is that metallic ions such as iron, manganese, or 

aluminum may be dissolved in concentrations which are 

subsequently directly toxic to plants. Under alkaline conditions, 

the danger to plants is the toxicity of sodium carbonates and 

bicarbonates either directly or indirectly (NAS, 1974). 

To avoid undesirable effects in irrigation waters, the pH 

should not exceed a range of 4.5 to 9.0. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
0 CRITERIA: 

PHENOL 

Aquatic Life 

The available data for phenol indicate that acute and chronic 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 10,200 and 2,560 ug/L, respectively, and would occur at 

lower concentrations among species that are more sensitive than 

those tested. 

The available data for phenol indicate that toxicity to 

saltwater aquatic life occurs at concentrations as low as 5,800 

ug/L and would occur at lower concentrations among species that 

are more sensitive than those tested. NO data are available 

concerning the chronic toxicity of phenol to sensitive saltwater 

aquatic life. 0 Human Health 

For comparison purposes, two approaches were used to derive 

criterion levels for phenol. Based on available toxicity data, 

to protect public health the derived level is 3.5 mg/L. 

Using available organoleptic data, to control 

undesirable taste and odor qualities of ambient water the 

estimated level is 0.3 mg/L. It should be recognized that 

organoleptic data have limitations as a basis for establishing a 

water quality criterion, and have no demonstrated relationship to 

potential adverse human health effects. 

NOTE: The U.S. EPA is currently developing Acceptable Daily 

Intake (ADI) or Verified Reference Dose (RfD) values for 

Agency-wide use for this chemical. The new value should 

be substituted when it becomes available. The January, 

1986, draft Verified Reference Dose document cites an RfD 

of 0.1 mg/kg/day for phenol. .'/ 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 

0 
PHOSPHORUS 

CRITERION:: 

0.10 ug/L yellow (elemental) phosphorus for marine or 

estuarine water. 

INTRODUCTION: 

Phosphorus in the elemental form is particularly toxic and is 

subject to bioaccumulation in much the same way as mercury. 

Phosphorus as phosphate is one bf the major nutrients required 

for plant nutrition and is essential for life. In excess of a 

cr it ica 1 concent rat ion, phosphates stirnu late p 1 ant growths. 

During the past 30 years, a formidable case has developed for the 

belief that increasing standing crops of aquatic plants, which 

often interfere with water uses and are nuisances to man, 

frequently are caused by increasing supplies of phosphorus. Such 

phenomena are associated with a condition of accelerated 

eutrophication or aging of waters. Generally, it is recognized 

that phosphorus is not the sole cause of eutrophication but there 

is substantiating evidence that frequently it is the key element 

of all of the elements required by freshwater plants, and 

generally, it is present in the least amount relative to need. 

Therefore, an increase in phosphorus allows use of other already 

present nutrients for plant growth. Further, of all of the 

elements required for plant growth in the water environment, 

phosphorus is the most easily controlled by man. 

Large deposits of phosphate rock are found near the western 

shore of Central Florida, as well as in a number of other States. 

Deposits in Florida are found in the form of pebbles which vary 

0 . ., 
in size from fine sand to about the size of a human foot. These 

pebbles are embedded in a matrix of clay and sand. The 

phosphate rock beds lie within a few feet of the surface and 

mining is accomplished by using hydraulic water jets and a 

washing operation that separates the phosphates from waste 

materials. The process is similar to that of strip-mining. 

Florida, Idaho, Montana, North Carolina, South Carolina, 

Tennessee, Utah, Virginia, and Wyoming share phosphate mining 

activities. 

Phosphates enter waterways from several different sources. 

The human body excretes about one pound per year of phosphorus 

expressed as I1Pt1. The use of phosphate detergents and other 

domestic phosphates increases the per capita contribution to 

about 3.5 pounds per year of phosphorus as P. Some industries, 

such as potato processing, have wastewaters high in phosphates. 

Crop, forest, idle, and urban land contribute varying amounts of 

phosphorus-diffused sources in drainage to watercourses. This 

drainage may be surface runoff of rainfall, effluent from tile 

lines, or return flow from irrigation. Cattle feedlots, 

concentrations of domestic duck or wild duck populations, tree 

leaves, and fallout from the atmosphere all are contributing 

sources. 

Evidence indicates that: (1) high phosphorus concentrations 

are associated with accelerated eutrophication of waters, when 

other growth-promoting factors are present: (2) aquatic plant 

problems develop in reservoirs and other standing waters at 

phosphorus values lower than those critical in flowing streams: 

(3) reservoirs and lakes collect phosphates from influent streams 
and store a portion of them within consolidated sediments, thus 

serving as a phosphate sink; and (4) phosphorus concentrations 

critical to noxious plant growth vary and nuisance growths may 

result from a particular concentration of phosphate in one 

geographical area but not in another. The amount or percentage 

of inflowing nutrients that may be retained by a lake or 

reservoir is variable and will depend upon: (1) the nutrient 

loading to the lake or resevoir; (2) the volume of the euphotic 

zone; (3) the extent of biological activities; (4) the detention 

time within a lake basin or the time available for biological 

activities; 

the penstock from the reservoir. 

and (5) the level of discharge from the lake or of 

Once nutrients are combined within the aquatic ecosystem, 

their removal is tedious and expensive. Phosphates are used by 

algae and higher aquatic plants and may be stored in excess of 

use within the plant cell. With decomposition of the plant cell, 

some phosphorus may be released immediately through bacterial 

action €or recycling within the biotic community, while the 

remainder may be deposited with sediments. Much of the material 

that combines with the consolidated sediments within the lake 

bottom is bound permanently and will not be recycled into the 

s ys tern. 

0 

RATIONAm : 

Elemental Phosphorus 

Isom (1960) reported an LC50 of 0.105 mg/L at 48 hours and 

0.025 mg/L at 160 hours for bluegill sunfish, __ Lepomis _--- 

y. ~ macrochirus, exposed to ye1 low phosphorus in distil led water at 
~ ~ - ~~ __ 

- 

26 OC and pH 7. The 125-and 195-hour LC50's of yellow phosphorus 

to Atlantic cod, Gadus morhua, and Atlantic salmon, Salmo salar, 

smolts in continuous-exposure experiments were 1.89 and 0.79 

ug/L, respectively (Fletcher and Hoyle, 1972). No evidence of an 

incipient lethal level was observed since the lowest 

concentration of p4 tested was 0.79 ug/L. Salmon that were 

exposed to elemental phosphorus concentrations of 40 ug/L or less 

developed a distinct external red color and showed signs of 

extensive hemolysis. The predominant features of p4 poisoning in 

salmon were external redness, hemolysis, and reduced hematocrits. 

Following the opening of an elemental phosphorus production 

plant in Long Harbour, Placentia Bay, Newfoundland, divers 

observed dead fish upon the bottom throughout the Harbour (Peer, 

1972). Mortalities were confined to a water depth of less than 18 

meters. There was visual evidence of selective mortality among 

benthos. Live mussels were found within 300 meters of the 

effluent pipe, while all scallops within this area were dead. 

Fish will concentrate elemental phosphorus from water 

containing as little as 1 ug/L (Idler, 1969). In one set of 

experiments, a cod swimming in water containing 1 ug/L elemental 

phosphorus for 18 hours concentrated phosphorus to 50 ug/kg in 

muscle, 150 ug/kg in fatty tissue, and 25,000 ug/kg in the liver 

(Idler, 1969; Jangaard, 1970). The experimental findings showed 

that phosphorus is quite stable in the fish tissues. 
The criterion of 0.10 ug/L elemental phosphorus for marine or 

estuarine waters is .1 of demonstrated lethal levels to important 

marine organisms and of levels that have been found to result in 

significant bioaccumulation. 

0 
Phosphate Phosphorus 

Although a total phosphorus criterion to control nuisance 

aquatic growths is not presented, it is believed that the 

following rationale to support such a criterion, which currently 

is evolving, should be considered. 

Total phosphate phosphorus concentrations in excess of 100 

ug/L P may interfere with coagulation in water treatment plants. 

When such concentrations exceed 25 ug/L at the time of the spring 

turnover on a volume-weighted basis in lakes or reservoirs, they 

may occasionally stimulate excessive or nuisance growths of algae 

and other aquatic plants. Algal growths inpart' undesirable 

tastes and odors to water, interfere with water treatment, become 

aesthetically unpleasant, and alter the chemistry of the water 

0 supply. They contribute to the phenomenon of cultural 

eutrophication. 

To prevent the development of biological nuisances and to 

control accelerated or cultural eutrophication, total phosphates 

as phosphorus (P) should not exceed 50 ug/L in any stream at the 

point where it enters any lake or reservoir, nor 25 ug/L within 

the lake or reservoir. A desired goal for the prevention of 

plant nuisances in streams or other flowing waters not 

discharging directly to lakes or impoundments is 100 ug/L total P 

(Mackenthun, 1973) Most relatively uncontaminated lake 

districts axe known to have surface waters that contain from 10 

to 30 ug/L total phosphorus as P (Hutchinson, 1957). 

The majority of the Nation's eutrophication problems are 0 

associated with lakes or reservoirs and currently there are more , .. .. 
data to support the establishment of a limiting phosphorus level 

in those waters than in streams or rivers that do not directly 

impact such water. There are natural conditions, also, that 

would dictate the consideration of either a more or less 

stringent phosphorus level. Eutrophication problems may occur in 

waters where the phosphorus concentration is less than that 

indicated above and, obviously, such waters would need more 

stringent nutrient limits. Likewise, there are those waters 

within the Nation where phosphorus is not now a limiting nutrient 

and where the need for phosphorus limits is substantially 

diminished. Such conditions are described in the last paragraph 

of this rationale. 

There are two basic needs in establishing a phosphorus 

criterion for flowing waters: ode is to control the development 

of plant nuisances within the flowing water and, in turn, to 

control and prevent animal pests that may become associated with 

such plants; the other is to protect the downstream receiving 

waterway, regardless of its proximity in linear distance. It is 

evidentthata portion ofthat phosphorus that enters a streamor 

other flowing waterway eventually will reach a receiving lake or 

estuary either as a component of the fluid mass, as bed load 

sediments that are carried downstream, or as floating organic 

materials that may drift just above the stream's bed or float on 

its water's surface. Superimposed on the loading from the 

inflowing waterway, a lake or estuary may receive additional 

phosphorus as fallout from the air shed or as a direct 

introduction from shoreline areas. 
Another method to control the inflow of nutrients, 

particularly phosphates, into a lake is that of prescribing an 

annual loading to the receiving water. Vollenweider (1973) 

suggests total phosphorus (P) loadings in grams per square meter 

of surface area per year that will be a critical level for 

eutrophic conditions within the receiving waterway for a 

particular water volume where the mean depth of the lake in 

meters is divided by the hydraulic detention time in years. 

Vollenweider's data suggest a range of loading values that should 

result in oligotrophic lake water quality. 

0 

Eutrophic 

or Critical 

Oligotrophic or 

Mean Depth/Hydraulic Permissible 

Detention Time Loading Loading 

(meters/year) (grams/meter2/year) (grams/meter2/year) 

0.5 

1.0 

2.5 

5.0 

7.5 

10.0 

25.0 

50.0 

75.0 

100.0 

0.07 

0.10 

0.16 

0.22 

0.27 

0.32 

0.50 

0.71 

0.87 

1-00 

0-14 

0.20 

0.32 

0.45 

0.55 

0.63 

1.00 

1.41 

1.73 

2.00 

There may be waterways wherein higher concentrations or 

loadings of total phosphorus do not produce eutrophy, as well as 

those waterways wherein lower concentrations or loadings of total 

0 

L-7 
phosphorus may be associated with populations of nuisance 

organisms. Waters now containing less than the specified amounts 

of phosphorus should not be degraded by the introduction of 

additional phosphates. 

It should be recognized that a number of specific exceptions 

can occur to reduce the threat of phosphorus as a contributor to 

lake eutrophy: 1. Naturally occurring phenomena may limit the 

development of plant nuisances. 2. Technological or costeffective limitations may help control introduced pollutants. 3. 

Waters may be highly laden with natural silts or colors which 

reduce the penetration of sunlight needed €or plant 

photosynthesis. 4. Some waters morphometric features of steep 

banks, great depth, and substantial flows contribute to a history 

of no plant problems. Wateks may be managed primarily for 

waterfowl or other wildlife. 7. In some waters nutrient other 

than phosphorus is limiting to plant growth: the level and nature 

of such limiting nutrient would. not be expected to increase to an 

extent that would influence eutrophication. 6. In some waters 

phosphorus control cannot be sufficiently effective under present 

technology to make phosphorus the limiting nutrient. 

5. 

No national criterion is presented for phosphate phosphorus 

for the control of eutrophication. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
PHTHALATE ESTERS 

CRITERIA: 

Aquatic Life 

The available data for phthalate esters indicate that acute 

and chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 940 and 3 ug/L, respectively, and would 

occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for phthalate esters indicate that acute 

toxicity to saltwater aquatic life occurs at concentrations as 

low as 2,944 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of phthalate esters to 

sensitive saltwater aquatic life but toxicity to one species of 

algae occurs at concentrations as low as 3.4 ug/L. 

, 

0 

Human Health 

For the protection of human health from the toxic properties 

of dimethyl phthalate ingested through water and contaminated 

aquatic organisms, the ambient water criterion is determined to 

be 313 mg/L. 

For the protection of human health from the toxic properties 

of dimethyl phthalate ingested through contaminated aquatic 

organisms alone, the ambient water criterion is determined to be 

2.9 g/l. 
For the protection of human health from the toxic properties 

of diethyl phthalate ingested through water and contaminated 

aquatic organisms, the ambient water criterion is determined to 

be 350 mg/L. 

For the protection of human health from the toxic properties 

of diethyl phthalate ingested through contaminated aquatic 

organisms alone, the ambient water criterion is determined to be 

1.8 g/l. 

For the protection of human health from the toxic properties 

of dibutyl phthalate ingested through water and contaminated 

aquatic organisms, the ambient water criterion is determined to 

be 34 mg/L. 

For the protection of human health from the toxic properties 

of dibutyl phthalate ingested through contaminated aquatic 

organisms alone, the ambient water criterion is determined to be 

154 mg/L. 

For the protection of human health from the toxic properties 

of di-2-ethylhexyl phthalate ingested through water and 

contaminated aquatic organisms, the ambient water criterion is 

determined to be 15 mg/L. 

For the protection of human health from the toxic properties 

of di-2-ethylhexyl phthalate ingested through contaminated 

aquatic organisms alone, the ambient water criterion is 

determined to be 50 mg/L. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
POLYCHLORINATED BIPHENYLS 

Aquatic Life 

For polychlorinated biphenyls the criterion to protect 

freshwater aquatic life as derived using the Guidelines is 0.014 

ug/L as a 24-hour average. The concentration of 0.014 ug/L is 

probably too high because it is based on bioconcentration factors 

measured in laboratory studies, but field studies apparently 

produce factors at least 10 times higher for fishes. The 

available data indicate that acute toxicity to freshwater aquatic 

life probably will occur only at concentrations above 2.0 ug/L 

and that the 24-hour average should provide adequate protection 

against acute toxicity. 

For polychlorinated biphenyls the criterion to protect 0 saltwater aquatic life as derived using the Guidelines is 0.030 

ug/L as a 24-hour average. The concentration of 0.030 ug/L is 

probably too high because it is based on bioconcentration factors 

measured in laboratory studies, but field studies apparently 

produce factors at least 10 times higher for fishes. The 

available data indicate that acute toxicity to saltwater aquatic 

life probably will only occur at concentrations above 10 ug/L and 

that the 24-hour average criterion should provide adequate 

protection against acute toxicity. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to polychlorinated biphenyls 0 

.i”i through ingestion of contaminated water and contaminated aquatic 
organisms, the ambient water concentration should be zero, based 

on the nonthreshold assumption for this chemical. However, 

zero level may not be attainable at the present time. Therefore, 

the levels which may result in incremental increase of cancer 

risk over the lifetime are estimated at and 

The corresponding recommended criteria are 0.79 ng/L, 0.079 ng/L, 

and 0.0079 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 0.79 ng/L, 0.079 ng/L, and 0.0079 ng/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
POLYNUCLEAR AROMATIC HYDROCARBONS 

CRITERIA : 

Aquatic Life 

The limited freshwater data base available for polynuclear 

aromatic hydrocarbons, mostly from short-term bioconcentration 

studies with two compounds, does not permit a statement 

concerning acute or chronic toxicity. 

The available data for polynuclear aromatic hydrocarbons 

indicate that acute toxicity to saltwater aquatic life occurs at 

concentrations as low as 300 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity of 

polynuclear aromatic hydrocarbons to sensitive saltwater aquatic 

life. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to polynuclear aromatic 

hydrocarbons through ingestion of contaminated water and 

contaminated aquatic organisms, the ambient water concentration 

should be zero, based on the nonthreshold assumption for this 

chemical. However, zero level may not be attainable at the 

present time. Therefore, the levels which may result in 

incremental increase of cancer risk over the lifetime are 

estimated at lom6, and The corresponding recommended 

criteria are 28.0 ng/L, 2.8 ng/L, and 0.28 ng/L, respectively. 

If these estimates are made for consumption of aquatic organisms 

only, excludinq consumption of water, the levels are 311.0 nq/L. 
31.1 ng/L, and 3.11 ng/L, respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
SELENIUM 

CRITERIA: 

Aquatic Life 

For total recoverable inorganic selenite the criterion 

to protect freshwater aquatic life as derived using the 

Guidelines is 35 ug/L as a 24-hour average, and the concentration 

should not exceed 260 ug/L at any time. 

For total recoverable inorganic selenite the criterion to 

protect saltwater aquatic life as derived using the Guidelines is 

54 ug/L as a 24-hour average, and the concentration should not 

exceed 410 ug/L at any time. 

The available data for inorganic selenate indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 760 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of inorganic selenate 

to sensitive freshwater aquatic life. 

No data are available concerning the toxicity of inorganic 

selenate to saltwater aquatic life. 

Human Health 

The ambient water quality criterion for selenium is 

recommended to be identical to the existing water standard which 

is 10 ug/L. Analysis of the toxic effects data resulted in a 

calculated level which is protective of human health against the 

ingestion of contaminated water and contaminated aquatic 

(- organisms. The calculated value is comparable to the present 

0 
standard. For this reason a selective criterion based on 

exposure solely from consumption of 6.5 grams of aquatic 

organisms was not derived. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
SILVER 

Aquatic - Life 

For freshwater aquatic life the concentration (in 

ug/L) of total recoverable silver should not exceed the 

numerical value given by e(l.72 [ ln(hardness) 3 - 6.52) at any 

time. For example, at hardnesses of 50, 100, and 200 mg/L as 

CaC03, the concentration of total recoverable silver should not 

exceed 1.2, 4.1, and 13 ug/L, respectively, at any time. The 

available data indicate that chronic toxicity to freshwater 

aquatic life may occur at concentrations as low as 0.12 ug/L. 

For saltwater aquatic life the concentration of total 

recoverable silver should not exceed 2.3 ug/L at any time. No 

data are available concerning the chronic toxicity of silver to 

sensitive saltwater aquatic life. 

Human Health 

The ambient water quality criterion for silver is 

recommended to be identical to the existing water standard, 

which is 50 ug/L. Analysis of the toxic effects data resulted in 

a calculated level which is protective of human health against 

the ingestion of contaminated water and contaminated aquatic 

organisms. The calculated value is comparable to the present 

standard. For this reason a selective criterion based on 

exposure solely from consumption of 6.5 grams of aquatic 

a organisms was not derived. 

- 

'. ' (45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
SOLIDS (DISSOLVED) AND - SALINlTY 

CRITERION: 

250 mg/L for chlorides and sulfates 

in domestic water supplies (welfare). 

INTRODUCTION: 

Dissolved solids and total dissolved solids are terms 

generally associated with freshwater systems and consist of 

inorganic salts, small amounts of organic matter, and dissolved 

materials (Sawyer, 1960). The equivalent terminology in Standard 

Methods is filtrable residue (Standard Methods, 1971). Salinity 

is an oceanographic term, and although not precisely equivalent 

to the total dissolved salt content it is related to it (Capurro, 

1970). For most purposes, the terms total dissolved salt content 

and salinity are equivalent. The principal inorganic anions 

dissolved in water include the carbonates, chlorides, sulfates, 

and nitrates (principally in ground waters) ; the principal 

cations are sodium, potassium, calcium, and magnesium. 

RATIONALE : 

Excess dissolved solids are objectionable in drinking water 

because of possible physiological effects, unpalatable mineral 

tastes, and higher costs because of corrosion or the necessity 

for additional treatment. 

The physiological effects directly related to dissolved 

solids include laxative effects principally from sodium sulfate 

and magnesium sulfate and the adverse effect of sodium on certain 

patients afflicted with cardiac disease and women with toxemia 

associated with pregnancy. One study was made using data 

e ,. ~ 
collected from wells in North Dakota. Results from a 

questionnaire showed that with wells in which sulfates ranged 

from 1,000 to 1,500 mg/L, 62 percent of the respondents indicated 

laxative effects associated with consumption of the water. 

However, nearly one-quarter of the respondents to the 

questionnaire reported difficulties when concentrations ranged 

from 200 to 500 mg/L (Moore, 1952). To protect transients to an 

area, a sulfate level of 250 mg/L should afford reasonable 

protection from laxative effects. 

As indicated, sodium frequently is the principal component of 

dissolved solids. Persons on restricted sodium diets may have an 

intake restricted from 500 to 1,000 mg/day (Nat. Res. Coun., 

1954). That portion ingested in water must be compensated by 

reduced levels in food ingested so that the total does not exceed 

the allowable intake. Using certain assumptions of water intake 

(e.g., 2 liters of water consumed per day) and sodium content of 

food, it has been calculated that for very restricted sodium 

diets, 20 mg/L in water would be the maximum, while for 

moderately restricted diets, 270 mg/L would be maximum. Specific 

sodium levels for entire water supplies have not been recommended 

but various restricted sodium intakes are recommended because: 

(1) the general population is not adversely affected by sodium, 

but various restricted sodium intakes are recommended by 

physicians for a significant portion of the population, and (2) 

270 mg/L of sodium is representative of mineralized waters that 

may be aesthetically unacceptable, but many domestic water 

supplies exceed this level. Treatment for removal bf sodium in 
water supplies is costly (NAS, 1974). 

A study based on consumer surveys in 29 California water 

systems was made to measure the taste threshold of dissolved 

salts in water (Bruvold et al., 1969). Systems were selected to 

eliminate possible interferences from other taste-causing 

substances than dissolved salts. The study revealed that 

consumers rated waters with 319 to 397 mg/L dissolved solids as 

llexcellentlt while those with 1,283 to 1,333 mg/L dissolved solids 

were "unacceptable" depending on the rating system used. A *lgoodll 

rating was registered for dissolved solids less than 658 to 755 

mg/L. The 1962 PHS Drinking Water Standards recommended a 

maximum dissolved solids concentration of 500 mg/L unless more 

suitable supplies were unavailable. 

Specific constituents included in the dissolved solids in 

water may cause mineral tastes at lower concentrations than other 

constituents. Chloride ions have frequently been cited as having 

a low taste threshold in water. Data from Ricter and MacLean 

(1939) on a taste panel of 53 adults indicated that 61 mg/L NaCl 

was the median level for detecting a difference from distilled 

water. At a median concentration of 395 mg/L chloride a salty 

taste was distinguishable, although the range was from 120 to 

1,215 mg/L. Lockhart, @t al. 1955) evaluated the effect of 

chlorides on water used for brewing coffee indicated threshold 

concentrations for chloride ranging from 210 mg/L to 310 mg/L 

depending on the associated cation. These data indicate that a 

level of 250 mg/L chlorides is a reasonable maximum level to 0 \.p protect consumers of drinking water. 
The causation of corrosion and encrustation of metallic 

surfaces by water containing dissolved solids is well known. In 

water distribution systems corrosion is controlled by insulating 

dissimilar metal connections by nonmetallic materials, using pH 

control and corrosion inhibitors, or some form of galvanic or 

impressed electrical current systems (Lehmann, 1964). In 

household systems water piping, wastewater piping, water heaters, 

faucets, toilet flushing mechanisms, garbage grinders and both 

clothes and dishwashing machines incure damage. 

By using water with 1,150 mg/L dissolved solids as compared 

with 250 mg/L, service life was reduced from 70 percent for 

toilet flushing mechanisms to 30 percent for washing equipment. 

Such increased corrosion was calculated in 1968 to cost the 

consumer an additional $0.50 per 1,000 gallons used. 

All species of fish and other aquatic life must tolerate a 

range of dissolved solids concentrations in order to survive 

under natural conditions. Based on studies in Saskatchewan it 

has been indicated that several common freshwater species 

survived 10,000 mg/L dissolved solids, that whitefish and pikeperch survived 15,000 mg/L, but only the stickleback survived 

20,000 mg/L dissolved solids. It was concluded that lakes with 

dissolved solids in excess of 15,000 mg/L were unsuitable for 

most freshwater fishes (Rawson and Moore, 1944). The 1968 NTAC 

Report also recommended 

less than that caused 

chloride. 

maintaining osmotic pressure levels of 

by a 15,000 mg/L solution of sodium 
Marine fishes also exhibit variance in ability to tolerate 

salinity changes. However, fishkills in Laguna Madre off the 

Texas coast have occurred with salinities in the range of 75 to 

100 o/oo. Such concentrated seawater is caused by evaporation 

and lack of exchange with the Gulf of Mexico (Rounsafell and 

Everhart, 1953). 

Estuarine species of fish are tolerant of salinity changes 

ranging from fresh to brackish to seawater. Anadromous species 

likewise are tolerant although evidence indicates that the young 

cannot tolerate the change until the normal time of migration 

(Rounsefell and Everhart, 1953). Other aquatic species are more 

dependent on salinity for protection from predators or require 

certain minimal salinities for successful hatching of eggs. The 

oyster drill cannot tolerate salinities less than 12.5 o/oo, 

Therefore, estuarine segments containing salinities below about 

12.5 o/oo produce most of the seed oysters for planting 

(Rounsefell and Everhart, 1953). Based on similar examples, the 

1968 NTAC Report recommended that to protect fish and other 

marine animals no changes in hydrography or stream flow should be 

allowed that permanently change isohaline patterns in the estuary 

by more than 10 percent from natural variation. 

Many of the recommended game bird levels for dissolved solids 

concentrations in drinking water have been extrapolated from data 

collected on domestic species such as chickens. However, young 

ducklings were reported poisoned in Suisan Marsh by salt when 

maximum summer salinities varied from 0.55 to 1.74 o/oo with 

~ means as high as 1.26 o/oo (Griffith, 1963). 
Indirect effects of excess dissolved solids are primarily the 

elimination of desirable food plants and other habitat-forming 

plants. Rapid salinity changes cause plasmolysis of tender 

leaves and stems because of changes in osmotic pressure. The 

1968 NTAC Report recommended the following limits in salinity 

variation from natural to protect wildlife habitats: 

Natural Salinity Variation Permitted 

(O/OO) (o/oo) 

0 to 3.5 1 

3.5 to 13.5 2 

13.5 to 35 4 

Agricultural uses of water are also limited by excessive 

dissolved solids concentrations. Studies have indicated that 

chickens, swine, cattle, and sheep can survive on saline waters 

up to 15,000 mg/L of salts of sodium and calcium combined with 

bicarbonates, chlorides, and sulfates but only 10,000 mg/L of 

corresponding salts of potassium and magnesium. The approximate 

limit for highly alkaline waters containing sodium and calcium 

carbonates is 5,000 mg/L (NTAC, 1968). 

Irrigation use of water depends not only upon the osmotic 

effect of dissolved solids, but also on the ratio of the various 

cations present. In arid and semiarid areas general 

classification of salinity hazards has been prepared (NTAC, 1968) 

(see Table 9). 

Table 9.-Dissolved Solids Hazard for Irrigation Water (mg/L). 

water from which no detrimental effects will usually 

be noticed--------------------- 500 .. - 

.. ,, 
water which can have detrimental effects on sensi500-1,000 tive crops--------------------- 

water that may have adverse 

effects on many crops and 

requires careful managemerit practices----------------- 1,000-2,000 

water that can be used for 

tolerant plants on perme- able soils with careful 

management practices----------- 2,000- 5,000 

The amount of sodium and the percentage of sodium in relation 

to other cations are often important. In addition to 

contributing to osmotic pressure, sodium is toxic to certain 

plants, especially fruits, and frequently causes problems in soil 

structure, infiltration, and permeability rates (Agriculture 

Handbook #60, 1954). A high percentage of exchangeable sodium in 

soils containing clays that swell when wet can cause a soil 

condition adverse to water movement and plant growth. The 

exchangeable-sodium percentage (ESP) * is an index of the sodium 

status of soils. An ESP of 10 to 15 percent is considered 

excessive if a high percentage of swelling clay minerals is 

present (Agricultural Handbook #60, 1954). 

0 

For sensitive fruits, the tolerance for sodium for irrigation 

water is for a sodium adsorption ratio (SAR)** of about 4, 

whereas for general crops and forages a range of 8 to 18 is 

generally considered usable (NTAC, 1968). It is emphasized that 

application of these factors must be interpreted in relation to 

specific soil conditions existing in a given locale and therefore 

frequently requires field investigation. . Industrial requirements regarding the dissolved solids ,- 

content of raw waters is quite variable. Table 10 indicates 
Table 10.-Total Dissolved Solids Concentrations of Surface 

Waters That Have Been Used as Sources for 

Industrial Water Supplies 

Industry/Use Maximum Concentration 

(mgm 

Textile 150 

Pulp and Paper 1,080 

Chemical 2,500 

Petroleum 3,500 

Primary Metals . 1,500 

Boiler Make-up 35,000 
maximum values accepted by various industries for process 0 requirements (NAS, 1974). Since water of almost any dissolved 

solids concentration can be de-ionized to meet the most stringent 

requirements, the economics of such treatment are the 1 imiting 

factor for industry. 

*ESP = 100 [a + b(SAR)] 

1 [a + b(SAR)] 

where: a = intercept respresenting experimental 

error 

(ranges from -0.06 to 0.01) 

from 0.014 to 0.016) 

b =slope of regression line (ranges 

**SAR = sodium adsorption ratio = Na - 

[0.5(Ca + Mg)]"" 

SAR is expressed as milliequivalents 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 

0 
SOLIDS (SUSPENDED, SETTLEABLE) - AND TURBIDITY 

CRITERIA 

Freshwater fish and other aquatic life: 

Settleable and suspended solids should not reduce the depth 

of the compensation point for photosynthetic activity by more 

than 10 percent from the seasonally established norm for 

aquatic life. 

INTRODUCTION: 

The term lkuspended and settleable solidst1 is descriptive of 

the organic and inorganic particulate matter in water. The 

equivalent terminology used for solids in Standard Methods (APHA, 

1971) is total suspended matter for suspended solids, settleable 

matter for settleable solids, volatile suspended matter for 

volatile solids and fixed suspended matter for fixed suspended 

solids. The term ltsolidstl is used in this discussion because of 0 

its more common use in the water pollution control literature. 

RAT1 ONALE : 

Suspended solids and turbidity are important parameters in 

b o t h m u n i c i p a l and industrial water supply practices. Finished 

drinking waters have a maximum limit of 1 turbidity unit where 

the water enters the distribution system. This limit is based on 

health considerations as it relates to effective chlorine 

disinfection. Suspended matter provides areas where 

microorganisms do not come into contact with the chlorine 

disinfectant (NAS, 1974). The ability of common water treatment 

processes (i.e., coagulation, sedimentation, filtration, and 

chlorination) to remove suspended matter to achieve acceptable 

final turbidities is a function of the composition of the 

material as well as its concentration. Because of the variability 

0 
of such removal efficiency, it is not possible to delineate a 

general raw water criterion for these uses. 

Turbid water interferes with recreational use and aesthetic 

enjoyment of water. Turbid waters can be dangerous for swimming, 

especially if diving facilities are provided, because ofthe 

possibility of unseen submerged hazards and the difficulty in 

locating swimmers in danger of drowning (NAS, 1974). The less 

turbid the water the more desirable it becomes for swimming and 

other water contact sports. Other recreational pursuits such as 

boating and fishing will be adequately protected by suspended 

solids criteria developed for protection of fish and other 

aquatic life. 

Fish and other aquatic life requirements concerning suspended 

solids can be divided into those whose effect occurs in the water 

column and those whose effect occurs following sedimentation to 

the bottom of the water body. Noted effects are similar for both 

fresh and marine waters. 

The effects of suspended solids on fish have been reviewed by 

the European Inland Fisheries Advisory Commission (EIFAC, 1965). 

This review in 1965 identified four effects on the fish and fish 

food populations, namely: 

(1) by acting directly on the fish swimming in water in which 

solids are suspended, and either killing them or reducing 

their growth rate, resistance to disease, etc.; 

(2) by preventing the successful development of fish eggs and 

larvae; 

(3) by modifying natural movements and migrations of fish; 
(4) by reducing the abundance of food available to the 

fish;. . . 

Settleable materials which blanket the bottom of water bodies 

damage the invertebrate populations, block gravel spawning beds, 

and if organic, remove dissolved oxygen from overlying waters 

(ElFAC, 1965; Edberg and Hofsten, 1973). In a study downstream 

from the discharge of a rock quarry where inert suspended solids 

were increased to 80 mg/L, the density of macroinvertebrates 

decreased by 60 percent while in areas of sediment accumulation 

benthic invertebrate populations also decreased by 60 percent 

regardless of the suspended sol id concantrations (Gammon, 1970). 

similar effects have been reported downstream from an area which 

was intensively logged. Major increases in stream suspended 

solids (25 ppm turbidity upstream versus 390 ppm downstream) 

caused smothering of bottom invertebrates, reducing organism 

density to only 7.3 per square foot versus 25.5 per square foot 

upstraam (Tebo, 1955). 

When settleable solids block gravel spawning beds which 

contain eggs, high mortalities result although there is evidence 

that some species of salmonids will not spawn in such areas 

(EIFAC, 1965). 

It has been postulated that silt attached to the eggs 

prevents sufficient exchange of oxygen and carbon dioxide between 

the egg and the overlying water. The important variables are 

particle size, stream velocity, and degree of turbulence (EIFAC, 

1965). 
Deposition of organic materials to the bottom sediments can 

cause imbalances in stream biota by increasing bottom animal 

density principally worm populations, and diversity is reduced as 

pol lution-sensitive forms disappear (Mackenthun, 1973). Algae 

1 ikewise flourish in such nutrient-rich areas although forms may 

become less desirable (Tarzwell and Gaufin, 1953). 

Plankton and inorganic suspended materials reduce light 

penetration into the water body, reducing the depth of thephotic 

zone. This reduces primary production and decreases fish food. 

The NAS commitee in 1974 recommended that the depth of light 

penetration not be reduced by more than 10 percent (NAS, 1974). 

Additionally, the near surface waters are heated because of the 

greater heat absorbency of the particulate material which tends 

to stabilize the water column and prevents vertical mixing (NAS, 

1974). Such mixing reductions decrease the dispersion of 

dissolved oxygen and nutrients to lower portions of the water 

body. 

One partially offsetting benefit of suspended inorganic 

material in water is the sorption of organic materials such as 

pesticides. Following this sorption process subsequent 

sedimentation may remove these materials from the water column 

into the sediments (NAS, 1974). 

Identifiable effects of suspended solids on irrigation use of 

water include the formation of crusts on top of the soil which 

inhibits water infiltration and plant emergence, and impedes soil 

aeration; the formation of films on plant leaves which blocks 

sunlight and impedes photosynthesis and which may reduce the 
marketability of some leafy crops like lettuce, and finally the 

adverse effect on irrigation reservoir capacity, delivery canals, 

and other distribution equipment (NAS, 1974). 

0 

The criterion for freshwater fish and other aquatic lifeare 

essentially that proposed by the National Academy of Sciences and 

the Great Lakes Water Quality Board. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
CRITERION: 

SULFIDE = HYDROGEN SULFIDE 

2 ug/L'undissociated HZS for 

fish and other aquatic life, fresh 

and marine water. 

INTRODUCTION: 

Hydrogen sulfide is a soluble, highly poisonous, gaseous 

compound havinq t h e c h a r a c t e r i s t i c odor of rotten eggs. It is 

detectable in air by humans at a dilution of 0.002 ppm. It will 

dissolve in water at 4,000 mg/L at 20' C and one atmosphere of 

pressure. Hydrogen sulfide biologically is an active compound 

that is found primarily as an anaerobic degradation product of 

both organic sulfur compounds and inorganic sulfates. Sulfides 

are constituents of many industrial wastes such as those from 

tanneries, paper mills, chemical plants, and gas works. The 

anaerobic decomposition of sewage, sludge beds, algae, and other 

naturally deposited organic material is a major source of 

hydrogen sulfide. 

0 

When soluble sulfides are added to water they react with 

hydrogen ions to form HS or HZS, the proportion of each depending 

on the pH. The toxicity of sulfides derives primarily from H2S 

rather than from the hydrosulfide (HS-) or sulfide (S=) ions' 

When hydrogen sulfide dissolves in water it dissociates according 

to the reactions: 

H2S HS- + H+ and HS- S= + H+ 

At pH 9 about 99 percent of the sulfide is in the form of HS- 

, at pH 7 the sulfide is equally divided between HS- and.H2S: and 0 

at pH 5 about 99 percent of the sulfide is present as H2S (NAS 

- 
_- 

a 1974). The fact that H2S is oxidized in well- aerated water 

by natural biological systems to sulfates or is biologically 

oxidized to elemental sulfur has caused investigators to minimize 

the toxic effects of H2S on fish and other aquatic life. 

RATIONALE: 

The degree of hazard exhibited by sulfide to aquatic animal 

life is dependent on the temperature, pH, and dissolved oxygen. 

At lower pH values a greater proportion is in the form of the 

toxic undissociated H2S. In winter when the pH is neutral or 

below or when dissolved oxygen levels are low but not lethal to 

fish, the hazard from sulfides is exacerbated. Fish exhibit a 

strong avoidance reaction to sulfide. Based on data from 

experiments with the stickleback, Jones (1964) hypothesized that 

if fish encounter a lethal concentration of sulfide there is a 

reasonable chance they will be repelled by it before they are 

harmed. This, of course, assumes that an escape route is open. 

Many past data on the toxicity of hydrogen sulfide to fish 

and other aquatic life have been based on extremely short 

exposure periods. Consequently, these early data have indicated 

that concentrations between 0.3 and 0.4 mg/L permit fish to 

survive (Van Horn 1958, Boon and Follis 1967, Theede et al., 

1969). Recent :ong-term data, both in field situations and under 

controlled 1aborator.y conditions, demonstrate hydrogen sulfide 

toxicity at lower concentrations. 

Colby and Smiti-i (1967) found that concentrations as high as 

0.7 mg/L have been found within 20 mm of the bottom of sludge 

beds, and the levels of 0.1 to 0.02 mg/L were common within the 
first 20 mm of water above this layer. Walleye (Stizostedion 

vitreum) eggs held in trays in this zone did not hatch. Adelman 

and Smith (1970) reported that the hatchability of northern pike 

(Esox lucius) eggs was substantially reduced at 25 ug/L H2S: at 

41 ug/L mortality was almost complete. Northern pike fry had 96- 

hour LC50 values that varied from 17 to 32 ug/L at normal oxygen 

levels of 6.0 mg/L. The highest concentration of hydrogen 

sulfide that had no observable effect on eggs and fry was 14 and 

4 ug/L, respectively. Smith and Oseid (1972), working on eggs, 

fry and juveniles of walleyes and white suckers (Catostomus 

commersoni) and Smith (1971), Safe levels in working on walleyes 

and fathead minnows, Pimephales promelas, were found to vary from 

2.9 ug/L to 12 ug/L with eggs being the least sensitive and 

juveniles being the most sensitive in short-term tests. In 96- 

hour bioassays, fathead minnows and goldfish, Carassius auratus, 

varied greatly in tolerance to hydrogen sulfide with changes in 

temperature. They were more tolerant at low temperatures (6 to 

10, C). Holland, et al. (1960) reported that 1.0 mg/L sulfide 

caused 100 percent mortality in 72 hours with Pacific salmon. 

0 

On the basis of chronic tests evaluating growth and survival, 

the safe H2S level for bluegill (Lepomis macrochirus) juveniles 

and adults was 2 ug/L. Egg deposition in bluegills was reduced 

after 46 days in 1.4 ug/L H2S (Smith and Oseid, 1974). White 

sucker eggs were hatched at 15 ug/L, but juveniles showed growth 

reductions at 1 ug/L. Safe level for fathead minnows were 

between 2 and 3 ug/L. Studies showed that safe levels for 

Gammarus Pseudolimnaeus and Hexagenia - limbata were 2 and 15 ug/L, 

respectively (Oseid and Smith, 1974a, 197413). Some species 

0 c 
typical of normally stressed habitats, Asellus spp., were much 

more resistant (Oseid and Smith, 1974~). 

Sulfide criteria for domestic or livestock use have not 

been established because the unpleasant odor and taste would 

preclude such use at hazardous concentrations. 

It is recognized that the hazard from hydrogen sulfide to 

aquatic life is often localized and transient. Available data 

indicate that water containing concentrations of 2.0 ug/L 

undissociated H2S would not be hazardous to most fish and other 

aquatic wildlife, but concentrations in excess of 2.0 ug/L would 

constitute a long-term hazard. 

I 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
TAINTING SUBSTANCES 

Materials should not be present in concentrations that 

individually or in combination produce undesirable flavors 

which are detectable by organoleptic tests performed on the 

edible portions of aquatic organisms. 

RATIONALE : 

Fish or shellfish with abnormal flavors, colors, tastes or 

odors are either not marketable or will result in consumer 

complaints and possible rejection of the food source even though 

subsequent lots of organisms may be acceptable. Poor product 

quality can and has seriously affected or eliminated the 

commercial fishing industry in some areas. Recreational fishing 

also can be affected adversely by off-flavored fish. For the 

majority of sport fishermen, the consumption of their catch is 

part of their recreation and off-flavored catches can result in 

diversion of the sportsmen to other water bodies. This can have 

serious economic impact on the established recreation industries 

such as tackle and bait sales and boat and cottage rental. 

0 

Water Quality Criteria, 1972 (NAS, 1974) lists a number of 

wastewaters and chemical compounds that have been found to lower 

the palatability of fish flesh. Implicated wastewaters included 

those from 2,4-D manufacturing plants, kraft and neutral sulfite 

pulping processes, municipal wastewater treatment plants, oily 

wastes, refinery wastes, phenolic wastes, and wastes from 

slaughterhouses. The 9 ist of imp1 icated chemical compounds is 

long: it includes cresol and phenol compounds, kerosene, 

naphthol, styrene, toluene, and exhaust outboard motor fuel. As 

little as 0.1 ug/L o-chlorophenol was reported to cause tainting 

a .. 
of fish flesh. 

Shumway and 

a 

Palensky 1973) determined estimated threshold 

concentrations for 22 organic compounds. The values ranged from 

0.4 ug/L (2 , 4-dichl orophenol ) to 9 5,000 ug/L (formaldehyde) . An 

additional 12 compounds were tested, 7 of which were not found 

to impair flavor at or near lethal levels. 

Thomas (1973) reviewed the literature review on tainting 

substances revealed serious problems that have occurred. Detailed 

studies and methodology used to evaluate the palatability of 

fishes in the Ohio River as affected by various waste discharges 

showed that the susceptibility of fishes to the accumulation of 

tainting substances is variable and dependent upon the species, 

length of exposure, and the polJutant. As little as 5 ug/L of 

gasoline can impart off-flavors to fish (Boyle, 1967). 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
TEMPERATURE 

Freshwater Aquatic Life 

For any time of year, there are two upper limiting 

temperatures for a location (based on the important sensitive 

species found there at that time): 

1. One limit consists of a maximum temperature for short 

exposures that is time dependent and is given by the speciesspecific equation: 

Temperature = (l/b) (log [time 3 -a) - 2, C 

(C,) 10 (min) 

where: loglo = logarithm to base 10 (common logarithm) 

a = intercept on the "y" or logarithmic axis 

of the l'ine fitted to experimental data 

and which is available for some species 

from Appendix 11-C, National Academy of 

Sciences 1974 document. 

b = slope of the line fitted to experimental 

data and available for some species from 

Appendix 11-C, of the National Academy 

of Sciences document. 

and 

2. The second value is a limit on the weekly average 

temperature that: 

a. In the cooler months (mid-October to mid-April in the 

north and December to February in the south) will 

protect against mortality of importr to mid-April in the 

north and December to February in the south) will 

protect against mortality of important species if the 

elevated plume temperature is suddenly dropped to the 

ambient temperature, with the limit being the 
b. 

acclimation temperature minus apt0 when the lower 

lethal threshold temperature equals the ambient water 

temperature (in some regions this limitation may also be 

applicable in summer). 

or 

In the warmer months (April through October in the north 

and March through November in the south) is determined 

by adding to the physiological optimum temperature 

(usually for growth) a factor calculated as one-third of 

the difference between the ultimate upper incipient 

lethal temperature and the optimum temperature €or the 

most sensitive important species (and appropriate life 

state) that normally is found at that location and time. 

or 

c. During reproductive seasons (generally April through 

June and September through October in the north and 

March through May and October through November in the 

south) the limit is that temperature that meets sitespecific requirements for successful migration, 

spawning, egg incubation, fry rearing, and other 

reproductive functions of important species. These 

local requirements should supersede all other 

requirements when they are applicable. 

or 

d. There is a site-specific limit that is found necessary 

to preserve normal species diversity or prevent 

appearance of nuisance organisms. 
Marine Aquatic - Life 

In order to assure protection of the characteristic 

indigenous marine community of a water body segment from adverse 

thermal effects: 

a. the maximum acceptable increase in the weekly 

average temperature resulting from artificial 

sources is 1' C (1.8 F) during all seasonsofthe 

year, providtng the summer maxima are not exceeded; 

and 

b. daily temperature cycles characteristic of the water 

body segment should not be altered in either 

amplitude or frequency. 

Summer thermal maxima, which define the upper thermal limits 

for the communities of the discharge area, should be established 

on a site-specific basis. Existing studies suggest the following 

regional limits: 

0 
Short-term Maximum 

Maximum True Daily Mean* 

Sub tropical regions (south of 

Cape Canaveral and Tampa Bay, 

Florida, and Hawaii 

32.2' C (90° F) 29.4O C (85' F) 

Cape Hatteras, N.C., to 

Cape Canaveral, Fla. 

Long Island (south shore) 30.6O C (87O F) 27.8O C (82O F) 

32.2' C (90' F) 29.4O C (85O'F) 

to Cape Hatteras, N.C. 

(* True Daily Mean = average of 24 hourly temperature readings.) 

Baseline thermal conditions should be measured at a site 

where there is no unnatural thermal addition from any source, 

which is in reasonable proximity to the thermal discharge (within 

5 miles) and which has similar hydrography to that of .the 

receiving waters at the discharge. 

INTRODUCTION: 

The uses of water by man in and out of its natural situs in 

the environment are affected by its temperature. Offstream 

domestic uses and instream recreation are both partially 

temperature dependent. Likewise, the 1 ife associated with the 

aquatic environment in any location has its species composition 

and activity regulated by water temperature. Since essentially 

all of these organisms are so-called "cold blooded" or 

poikilotherms, the temperature of the water regulates their 

metabolism and ability to survive and reproduce effectively. 

Industrial uses for process water and for coolingare likewise 

regulated by the water's temperature. Temperature, therefore, is 

an important physical parameter which to some extent regulates 

many of the beneficial uses of water. The Federal Water 

Pollution Control Administration in 1967 called temperature a 
catalyst, a depressant, an activator, a restrictor, a stimulator, 

a controller, a killer, one of the most important and most 

influential water quality characteristics to life in water." 

0 

RATIONALE : 

The suitability of water for total body immersion is greatly 

affected by temperature. In temperate climates, dangers from 

exposure to low temperatures is more prevalent than exposure to 

elevated water temperatures. Depending on the amount of activity 

by the swimmer, comfortable temperatures range from 20° C to 30° 

e. Short durations of lower and higher temperatures can be 

tolerated by most individuals. For example, for a 30-minute 

period, temperatures of 10' C or 35O C can be tolerated without 

harm by most individuals (NAS, 1974). 

Temperature also affects the self-purification phenomenon in 

water bodies and therefore the aesthetic and sanitary qualities 

that exist. Increased temperatures accelerate the biodegradation 

of organic material both in the overlying water and in bottom 

deposits which makes increased demands on the dissolved oxygen 

resources of a given system. The typical situation is exacerbated 

by the fact that oxygen becomes less soluble as water temperature 

increases. Thus, greater demands are exerted on an increasingly 

scarce resource which may lead to total oxygen depletion and 

obnoxious septic conditions. These effects have been described by 

Phelps (1944) , Carp (1963), and Velz (1970). 

Indicator enteric bacteria, and presumably enteric pathogens, 

are likewise affected by temperature. It has been shown that 

- both total and fecal coliform bacteria die away more rapidly in 

the environment with increasing temperatures (Ballentine and 
Kittrell, 1968). 

Temperature effects have been shown for water treatment 

processes. Lower temperatures reduce the effectiveness of 

coagulation with alum and subsequent rapid sand filtration. In 

one study, difficulty was especially pronounced below 5O C 

(Hannah, et al., 1967). Decreased temperature also decreases the 

effectiveness of chlorination. Based on studies relating 

chlorine dosage to temperature, and with a 30-minute contact 

time, dosages required for equivalent disinfective effect 

increased by as much as a factor of 3 when temperatures were 

decreased from 20° C to loo C (Reid and Carlson, 1974). 

Increased temperature may increase the odor of water because of 

the increased volatility of odor-causing compounds (Bumson, 

1938). Odor problems associated with plankton may also be 

aggravated. 

The effects of temperature on aquatic organisms have been the 

subject of comprehensive literature reviews (Brett, 1956; Fry, 

1967; FWPCA, 1967; Kine, 1970) and annual literature reviews 

published by the Water Pollution Control Federaticn (Coutant, 

1968, 1969, 1970, 1971; Coutant and Goodyear, 1972; Coutant and 

Pfuderer, 1973, 1974). Only highlights from the thermal effects 

on aquatic life are presented here. 

Temperature changes in water bodies can alter the existing 

aquatic community. The dominance of various phytoplankton groups 

in specific temperature ranges has been shown. For example, from 

20° C to 25' C, diatoms predominated; green algae predominated 

from 30' C: to 35O C and blue-greens predominated above 3.5' C 
airns, 1956). Likewise, changes from a coldwater fishery to a 

warm-water fishery can occur because temperature may be directly 

lethal to adults or fry cause a reduction of activity or limit 

0 (c 

reproduction (Brett, 1960) 

Upper and lower limits for temperature have been established 

for many aquatic organisms. Considerably more data exist for 

upper as opposed to lower limits. Tabulations of lethal 

temperatures for fish and other organisms are available (Jones, 

1964: FWPCA, 1967 NAS, 1974). Factors such as diet, activity, 

age, general health, osmotic stress, and even weather contribute 

to the lethality of temperature. The aquatic species, thermal 

accumulation state and exposure time are considered the critical 

factors (Parker and Xrenkel, 1969). 

The effects of sublethal temperatures on metabolism, 

respiration, behavior, distribution and migration, feeding rate, 

growth, and reproduction have been summarized by Be Sylva (1969). 

Another study has i l l u s t r a t e d t h a t inside the tolerance zone 

there is encompassed a more restrictive temperature range in 

which normal activity and growth occur and yet an even more 

restrictive zone inside that in which normal reproduction will 

occur (Brett, 1960). 

De Sylva (1969) has summarized available data on the combined 

effects of increased temperature and toxic materials on fish 

indicate that toxicity generally increases with increased 

temperature and that organisms subjected to stress from toxic 

materials are less tolerant of temperature extremes. 

The tolerance of organisms to extremes of temperature is a 

function of their genetic ability to adapt to thermal changes 

0 ~ 
within their characteristic temperature range, the acclimation 

temperature prior to exposure, and the time of exposure to the 

elevated temperature (Coutant, 1972). The upper incipient lethal 

temperature or the highest temperature that 50 percent of a 

sample of organisms can survive is determined on the organism at 

the highest sustainable acclimation temperature. The lowest 

temperature that 50 percent of the warm acclimated organisms can 

survive in is the ultimate lower incipient lethal temperature. 

True acclimation to changing temperatures requires several days 

(Brett, 1941). The lower end of the temperature accommodation 

range for aquatic life is 0' C in fresh water and somewhat less 

for saline waters. However, organisms acclimated to relatively 

warm water, when subjected to reduced temperatures that under 

other conditions of acclimation would not be detrimental, may 

suffer a significant mortality caused by thermal shock (Coutant, 

1972). 

Through the natural changes in climatic conditions, the 

temperatures of water bodies fluctuate daily, as well as 

seasonally. These changes do not eliminate indigenous aquatic 

populations, but affect the existing community structure and the 

geographic distribution of species. Such temperature changes are 

necessary to induce the reproductive cycles of aquatic organisms 

and to regulate other life factors (Mount, 1969). 

Artificially induced changes such as the return of cooling 

water or the release of cool hypolimnetic waters from 

impoundments may alter indigenous aquatic ecosystems (Coutant, 

1972). Entrained organisms may be damaged by temperature 
increases across cooling water condensers if the increase is 

sufficiently great or the exposure period sufficiently long. 

Impingement upon condenser screens, chlorination for slime 

control, or other physical insults damage aquatic life (Raney, 

1969: Patrick, 1969 (b)). However, Patrick (1969(a)) has shown 

that algae passing through condensers are not injured if the 

temperature of the outflowing water does not exceed 345O C. 

In open waters elevated temperatures nay affect periphyton, 

benthic invertebrates, and fish, in addition to causing shifts in 

algal dominance. Trembley (1960) studies of the Delaware River 

downstream from a power plant concluded that the periphyton 

population was considerably altered by the discharge. 

The number and distribution of bottom organisms decrease as 

water temperatures increase. The upper tolerance limit for a 0 

balanced benthic population structure is approximately 32O C, A 

large number of these invertebrate species are able to tolerate 

higher temperatures than those required for reproduction (FWPCA, 

1967). 

In order to define criteria for fresh waters, Coutant (1972) 

cited the following was cited as currently definable 

requirements: 

1. Maximum sustained temperatures that are consistent with 

maintaining desirable levels of productivity, 

2. maximum levels of metabolic acclimation to warm 

temperatures that will permit return to ambient winter 

temperatures should artificial sources of heat cease, 

3. Time-dependent temperature 1 imitations for survival of 

brief exposures to temperature extremes, both upper and lower, 
4. Restricted temperature ranges for various states of 

reproduction, including (for fish) gametogenesis, spawning 

migration, release of gametes, development of the embryo, 

commencement of independent feeding (and other activities) by 

j uv eni 1 es , and temper a tur es re qu ired for met amorphos is, 

emergence, or other activities of lower forms, 

5. Thermal limits for diverse species compositions of aquatic 

communities, particularly where reduction in diversity 

creates nuisance growths of certain organisms, or where 

important food sources (food chains) are altered, 

6. Thermal requirements of downstream aquatic life (in 

rivers) where upstream diminution of a coldwater resource 

will adversely affect downstream temperature requirements. 

The major portion of such information that is available, 

however, is for freshwater fish species rather than lower forms 

of marine aquatic life. 

The temperature-time duration for short-term exposures such 

that 50 percent of a given population will survive an extreme 

temperature frequently is expressed mathematically by fitting 

experimental data with a staright line on a semi-logarithmic plot 

with time on the logarithmic scale and temperature on the linear 

scale (see fig. 1). In equation form this 50 percent mortality 

relationship is: 

loglo (time (minutes)) = a + b (Temperature (O C)) 

where: loglo= logarithm to base 10 (common logarithm) 

a = intercept on the "y" or logarithmic axis of 

the line fitted to experimental data and which 

is available for some species from Appendix 11-C, 

of the National Academy of Sciences document. 

b = slope of the line fitted to experimental data 

and which is available for some species from 

Appendix 11-C, of the National Academy of 

Sciences document. 

To provide a safety factor so that none or only a few 

organisms will perish, it has been found experimentally that a 
criterion of 2O C below maximum temperature is usually sufficient 

(Black, 1953). To provide safety for all the organisms, the 

temperature causing a median mortality for 50 percent of the 

population would be calculated and reduced by 2' C in the case 

of an elevated temperature. Available scientific information 

includes upper and lower incipient lethal temperatures, 

coefficients I1at1 and llbll for the thermal resistance equation, and 

information of size, life stage, and geographic source of the 

particular test species (Appendix 11-C, NAS, 1974). 

Maximum temperatures for an extensive exposure (e.g., more 

than 1 week) must be divided into those for warmer periods and 

winter. Other than for reproduction, the most temperaturesensitive life function appears to be growth (Coutant, 1972). 

Coutant (1972) has suggested that a satisfactory estimate of a 

limiting maximum weekly mean temperature may be an average of the 

optimum temperature for growth and the temperature €or zero net 

growth. 

Because of the difficulty in determining the temperature of 

zero net growth, essentially the same temperature can be derived 

by adding to the optimum essentially to temperature (for growth 

or other physiological functions) a factor calculated as onethird of the difference between the ultimate upper incipient 

lethal temperature and the optimum temperature (NAS, 1974). In 

equation form: 

Maximum weekly (ultimate upper optimum) 

average = optimum + 1/3 (incipient lethal - temperature) 

temperature temperature (temperature) 

Since temperature tolerance varies with various states of 

development of a particular species, the criterion for a 

- 
particular location would be calculated for the most important 

life form likely to be present during a particular month. One 

caveat in using the maximum weekly mean temperature is that the 

limit for short-term exposure must not be exceeded. Example 

calculations for predicting the summer maximum temperatures for 

short-term survival and for extensive exposure for various fish 

species are presented in Table 11. These calculations use the 

above equations and data from EPA's Environmental Research 

Laboxatory in Duluth. 

The winter maximum temperature must not exceed the ambient 

water temperature by more than the amount of change a specimen 

acclimated to the plume temperature can tolerate. Such a change 

could occur by a cessation of the source of heat or by the 

specimen being driven from an area by addition of biocides or 

other factors. However, there are inadequate data to estimate a 

safety factor for the Isno stress" level from cold shocks (NAS, 

1974). Figure 2 was developed from available data in the 

literature (ERL-Duluth, 1976) and can be used for estimating 

allowable winter temperature increases. 

Coutant (1972) has reviewed the effects of temperature on 

aquatic life reproduction and development. Reproductive events 

are noted as perhaps the most thermally restricted of all life 

phases assuming other factors are at or near optimum levels. 

Natural short-term temperature fluctuations appear to cause 

reduced reproduction of fish and invertebrates. 
TABLE 11.-Example Calculated Values for 

Maxima for Survival for Juveniles and 

Adults During the Summer 

(Centigrade and Fahrenheit). 

Species Growtha Maxima 

Maximum Weekly Average Temperatures for Growth and Short-Term 

b 

Atlantic salmon 

Bigmouth buffalo 

Black crappie 

Bluegill 

Brook trout 

Carp 

Channel catfish 

Coho salmon 

Emerald shiner 

Freshwater drum 

Lake herring (Cisco) 

Largemouth bass 

Northern pike 

Rainbow trout 

Sauger 

Smallmouth bass 

Smallmouth buffalo 

Sockeye salmon 

Striped bass 

Threadfin shad 

White bass 

White crappie 

White sucker 

Yellow perch 

20 (68) 

27 (81) 

32 (90) 

19 (66) 

32 (90) 

18 (64) 

30 (86) 

17 (63) 

32 (90) 

28 (82) 

19 (66) 

25 (77) 

29 (84) 

18 (64) 

28 (82) 

28 (82) 

29 (84) 

23 (73) 

35 (95) 

24 (75) 

35 (95) 

24 (75) 

25 (77) 

30 (86) 

24 (75) 

34 (93) 

22 (72) 

a - Calculated according to the equation (using optimum 

temperature for growth) 

maximum weekly average temperature for growth = optimum 

temperature + 1/3 (ultimate incipient lethal temperatureoptimum temperature. 

b - Based on temperature (OC) = l/b (log” time(min.) -a) 

2O C, acclimation at the maximum weekly average temperature 

€or summer growth, and data in Appendix 11-C of Water 

Quality Criteria, published by National Academy of Sciences. 0 

c - Based on data for larvae (ERL-Duluth, 1976). 

-. , 
There are indadequate data available quantitating the most 

temperature-sensitive life stages among various aquatic species. 

Uniform elevation of temperature a few degrees but still within 

the spawning range may lead to advanced spawning for spring 

spawning species and delays for fall spawners. Such changes may 

not be detrimental unless asynchrony occurs between newly 

hatched juveniles and their normal food source. Such asynchrony 

may be most pronounced among anadromous species or other migrants 

who pass from the warmed area to a normally chilled, unproductive 

area. Reported temperature data on maximum temperatures for 

spawning and embryo survival have been summarized in Table 12 

(from ERL-Duluth 1976). 

Although the limiting effects of thermal addition to 

estuarine and marine waters are not as conspicuous in the fall, 

winter, and spring as during the summer season of maximum heat 

stress, nonetheless crucial thermal limitations do exist. Hence, 

it is important that the thermal additions to the receiving 

waters be minimized during all seasons of the year. Size of 

harvestable stocks of commercial fish and shellfish, particularly 

near geographic limits of the fishery, appear to be markedly 

influenced by slight changes in the long-term temperature regime 

(Dow, 1973). 

Jefferies and Johnson (1974) studied the relationship between 

temperature and annual variation in 7-year catch data for winter 

flounder,Pseudopleuronectes _-__-_--_-I americanus in Narragansett Bay, 

Rhode Island, revealed that a 78 percent decrease in annual catch 

correlated closely with a 0.5OC increase in the average 
temperature over the 30-month period between spawning and 

recruitment into the fishery. Sissenwine's 1974 model predicts a 

68 percent reduction of recruitment in ye1 Powtail flounder, 

Limanda -ferrugiia, -- with a l0C long-term elevation in southern 

New England waters. 
TABLE 12. 

Summary of Reported Values for 

Maxima for Embryo Survival During the Spawning Season 

(Centigrade and Fahrenheit) 

Maximum Weekly Average Temperature for Spawning and Short-Term 

Species 

Atlantic Salmon 

Bigmouth Buffalo 

Black Crappie 

Bluegill 

Brook Trout 

carp 

Channel Catfish 

Coho Salmon 

Emerald Shiner 

Freshwater Drum 

Lake Herring (Cisco) 

Largemouth Bass 

Northern Pike 

Rainbow Trout 

Sauger 

Smallmouth Bass 

Smallmouth Buffalo 

Sockeye Salmon 

Striped Bass 

Threadfin Shad 

White Bass 

White Crappie 

White Sucker 

Yellow Perch 

Spawning, Embryo 

Survivalb 

5 

17 

25 

9 

21 

27 

10 

24 

21 

3 

21 

11 

9 

17 

17 

10 

18 

18 

17 

18 

10 

12 

I 10 

(41) 7 

77 1 34 

70) 33 

(63) 27 

48) 13 

81) 29 

50) 13 

75) 28 

70) 26 

37) 8 

70) 27 

52) 19 

48) 13 

50) 21 

63 1 

21 

13 

63) 

50) 

64 1 24 

64 1 34 

63) 26 

23 

20 

64) 

20 

50) 

54) 

a - the optimum or mean of the range of spawning temperatures 

reported for the species (ERL-Duluth, 1976). 

b - the upper temperature for successful incubation and 

hatching reported for the species (ERL-Duluth, 1976) - 

c - upper temperature for spawning. 
Community balance can be influenced strongly by such 

temperature-dependent factors as rates of reproduction, 

recruitment, and growth of each component population. A few 

degrees elevation in average monthly temperature can appreciably 

alter a community through changes in interspecies relationships. 

A 50 percent reduction in the softshell clam fishery in Maine by 

the green crab, Carcinus maenus, illustrates how an increase in 

winter temperatures can establish new predator-prey 

relationships. Over a period of 4 years, there was a natural 

amelioration of temperature and the monthly mean for the coldest 

month of each year did not fall below 2OC. This apparently 

precluded appreciable ice formation and winter cold kill of the 

green crab and permitted a major expansion of its population, 

with increased predation of the softshell clam resulting (Glude, 

1954: Welch, 1968). 

Temperature is a primary factor controlling reproduction and 

can influence many events of the reproductive cycle from 

gametogenesis to spawning. Among marine invertebrates, 

initiation of reproduction (gametogenesis) is often triggered 

during late winter by attainment of a minimum environmental 

threshold temperature. In some species, availability of adequate 

food is also a requisite (Pearse, 1970; Sastry, 1975: devlaming, 

1971). Elevated temperature can limit gametogenesis by 

preventing accumulation of nutrients in the gonads. This problem 

could be acute during the winter if food availability and feeding 

activity is reduced. Most marine organisms spawn during the 0 spring and summer; gametogenesis is usually initiated during the 
previous fall. It should also be noted that some species spawn 

only during the fall (herrinhg) ,while others during the winter 

and very early spring. At the higher latitudes, winter breeders 

include such estuarine community dominants as acorn barnacles, 

-- Balanus - balanus and B. balanoides, the edible blue mussel Mytilus _ 

-edulis ---I sea urchin, Strongylocentrotus drobachiensis, sculpin, 

and the winter flounder, Pseudopleuronectes - americanus. The two 

boreal barnacles require temperatures below 10°C before egg 

production will be initiated (Crisp, 1957). It is clear that 

adaptations for reproduction exist which are dependent on 

temperature conditions close to the natural cycle. 

Juvenile and adult fish usually thermoregulate behaviorally 

by moving to water having temperatures closest to their thermal 

preference. This provides a thermal environment which 

approximates the optimal temperature for many physiological 

functions, including growth (Neil1 and Magnuson. 1974). As a 

consequence, fishes usually are attracted to heated water during 

the fall, winter, and spring. Avoidance will occur as warmer 

temperature exceeds the preferendum by 1 to 3OC (Coutant, 1975). 

This response precludes problems of heat stress for juvenile and 

adult fishes during the summer, but several potential problems 

exist during the other seasons. The possibility of cold shock 

and death of plume-entrained fish resulting from winter plant 

shutdown is well recognized. Also, increased incidence of 

disease and a deterioration of physiological condition has been 

observed among plume-entrained fishes, perhaps because of 

insufficient food (Massengill, 1973). A weight loss of 

approximately 10 percent for each lo C rise in water temperature 
has been observed in fish when food is absent. (Phillips et al., 

1960) There may also be indirect adverse effects on the 

indigenous community because of increased predation pressure if 

thermal addition leads to a concentration of fish which are 

dependent on this community for their food. 

Fish migration is often linked to natural environmental 

temperature cycles. In early spring, fish employ temperature as 

their environmental cue to migrate northward (e.g., menhaden, 

bluefish) or to move inshore (winter flounder). Likewise, water 

temperature strongly influences timing of spawning runs ofanadromous fish into rivers (Leggett and Whitney, 1972). In the 

autumn, a number of juvenile marine fishes and shrimp are 

dependent on a drop in temperature to trigger their migration 

from estuarine nursery grounds for oceanic dispersal or southward 

migration (Lund and Maltezos, 1970; Talbot, 1966). 

Thermal discharges should not alter diurnal and tidal 

temperature variations normally experienced by marine 

communities. Laboratory studies show thermal tolerance to be 

enhanced when animals are maintained under a diurnally 

fluctuating temperature regime rather than at a constant 

temperature (Costlow and Bookhout, 1971; Furch, 1972; Hoss, et 

al.,). A daily cyclic regime can be protective additionally as 

it reduces duration of exposure to extreme temperatures (Pearce, 

1969; Gonzalez, 1972). 

0 

Summer thermal maxima should be established to protect the 

various marine communities within each biogeographic region. 

During the summer, naturally elevated temperatures may be of -1 
sufficent magnitude to cause death or emigration (Glynn, 1968; 

Vaughn, 1961). This more commonly occurs in tropical and warm 

temperate zone waters, but has been reported for enclosed bays 

and shallow waters in other regions as well (Nichols, 1918). 

Summer heat stress also can contribute to increased incidence of 

disease or parasitism (Sinderman, 1965) : reduce or block sexual 

maturation (Thorhaug, et al., 1971: deVlaming, 1972); inhibit or 

block embryonic cleavage of larval development (Calabrese, 1969) ; 

reduce feeding and growth of juveniles and adults (011a and 

Studholme, 1971) : result in increased predation (Gonzalez, 1972); 

and reduce productivity of macroalgae and seagrasses (South and 

Hill, 1970; Zieman, 1970). The general ceilings set forth here 

are derived from studies delineating limiting temperatures for 

the more thermally sensitive species or communities of a 

biogeographic region. 

Thermal effects data are presently insufficient to set 

general temperature limits for all coastal biogeographic regions. 

The data enumerated in the Appendix, plus any additional data 

subsequently generated, should be used to develop thermal limits 

which specifically consider communities relevant to given water 

bodies. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOLOGY 
2,3,7,8-TETRACHMRODIBENZO-P-DIOXIN 

CRITERIA: 

Aquatic Life 

Not enough data are available concerning the effects of 

2,3,7,8-TCDD on aquatic life and its uses to allow derivation of 

national criteria. The available information indicates that 

acute values for some freshwater animal species are >1.0 ug/L; 

some chronic values are <0.01 ug/L; and the chronic value 

for rainbow trout is <0.001 ug/L. Because exposures of 

some species of fishes to 0.01 ug/L for <6 days resulted in 

substantial mortality several weeks later, derivation of 

aquatic life criteria for 2,3,7,8-TCDD may require special 

consideration. Predicted bioconcentration factors (BCFs) for 

2,3,7,8-TCDD range from 3,000 to 900,000, but the available 0 measured BCFs range from 390 to 13,000. If the BCF is 5,000, 

concentrations >0.00001 ug/L should result in concentrations 

in edible freshwater and saltwater fish and shellfish that 

exceed levels identified in a U.S. FDA health advisory. If the 

BCF is >5,000 or if uptake in a field situation is greater than 

that in laboratory tests, the value of 0.00001 ug/L will be too 

high. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of 2,3,7,8-TCDD exposure through ingestion 

of contaminated water and contaminated aquatic - orqanisms, . the 

ambient water concentration should be zero. This criterion is 
based on the nonthreshold assumption for 2,3,7,8-TCDD. However, 

zero may not be an attainable level at this time. 

(49 F.R. 5831, February 15, 1984) 

SEE APPENDIX B FOR METHODOLOGY 
TETRACHLOROETHYLENE 

Aquatic Life 

The available data for tetrachloroethylene indicate 

that acute and chronic toxicity to freshwater aquatic life 

occurs at concentrations as low as 5,280 and 840 ug/L, 

respectively, and would occur at lower concentrations among 

species that are more sensitive than those tested. 

The available data for tetrachloroethylene indicate 

that acute and chronic toxicity to saltwater aquatic life 

occurs at concentrations as low as 10,200 and 450 ug/L, 

respectively, and would occur at lower concentrations among 

species that are more sensitive than those tested. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to tetrachloroethylene through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, the 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at and 

The corresponding recommended criteria are 8.0 ug/L, 0.80 ug/L, 

and 0.08 ug/L, respectively. If these estimates are made for 
consumption of aquatic organisms only, excluding consumption of 

water, the levels are 88.5 ug/L, 8.85 ug/L, and 0.88 ug/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERIA: 

THALLIUM 

Aquatic Life 

The available data for thallium indicate that acute and 

chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 1,400 and 40 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. Toxicity to one species of fish 

occurs at concentrations as low as 20 ug/L after 2,600 hours of 

exposure. 

The available data for thallium indicate that acute toxicity 

to saltwater aquatic life occurs at concentrations as low as 

2,130 ug/L and would occur at lower concentrations among species 

that are more sensitive than those tested. No data are 0 available concerning the chronic toxicity of thallium to 

sensitive saltwater aquatic life. 

Human Health 

For the protection of human health from the toxic properties 

of thallium ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 13 

U9/L. 

For the protection of human health from the toxic properties 

of thallium ingested through contaminated aquatic organisms 

alone, the ambient water criterion is determined to be 48 ug/L. 

0 (45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
m CRITERIA: 

TOLUENE 

Aquatic Life 

The availab-2 data for toluene indicate t at acute toxicity 

to freshwater aquatic life occurs at concentrations as low as 

17,500 ug/L and would occur at lower concentrations among species 

that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of toluene to 

sensitive freshwater aquatic life. 

The available data for toluene indicate that acute and 

chronic toxicity to saltwater aquatic life occurs at 

concentrations as low as 6,300 and 5,000 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 0 Human Health 

For the protection of human health from the toxic properties 

of toluene ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 14.3 

w/L. 

For the protection of human health from the toxic properties 

of toluene ingested through contaminated aquatic organisms 

alone, the ambient water criterion is determined to be 424 

W/L- 

(45 F.R. 79318, November 28, 1980) 

NOTE: The U.S. EPA is currently developing Acceptable Daily 

Intake (ADI) or Verified Reference Dose (RfD) values €or 

Agency-wide use for this chemical. The new value should 

be substituted when it becomes available. The January, * - 1986, draft Verified Reference Dose document cites an RfD 

of 0.3 mg/kg/day for toluene. 

SEE APPENDIX B FOR METHODOLOGY 

0 
TOXAPHENE 

CRITERIA: 

Aquatic - Life 

For toxaphene the criterion to protect freshwater aquatic 

life as derived using the Guidelines is 0.013 ug/L as a 24-hour 

average, and the concentration should not exceed 1.6 ug/L at any 

time. 

For saltwater aquatic life the concentration of 

toxaphene should not exceed 0.070 ug/L at any time. No data are 

available concerning the chronic toxicity of toxaphene to 

sensitive saltwater aquatic life. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to toxaphene through ingestion 

of contaminated water and contaminated aquatic organisms, 

the ambient water concentration should be zero, based on the non 

threshold assumption for this chemical. However, zero level may 

not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

lifetime are estimated at and The 

corresponding recommended criteria are 7.1 ng/L, 0.71 ng/L, and 

0.07 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 7.3 ng/L, 0.73 ng/L, and 0.01 ng/L, 

0 respectively. 

-. 

\- 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERIA: 

TRICHLOROETHYLENE 

Aquatic Life 

The available data for trichloroethylene indicate that acute 

toxicity to freshwater aquatic life occurs at concentrations as 

low as 45,000 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of trichloroethylene to 

sensitive freshwater aquatic life but the behavior of one species 

is adversely affected at concentrations as low as 21,900 ug/L. 

The available data for trichloroethylene indicate that acute 

toxicity to saltwater aquatic life occurs at concentrations as 

low as 2,000 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of trichloroethylene to 

sensitive saltwater aquatic life. 

0 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to trichloroethylene through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentration should be zero, based 

on the nonthreshold assumption for this chemical. However, 

zero level may not be attainable at the present time. Therefore, 

the levels which may result in incremental increase of cancer 

risk over the lifetime are estimated at lom5, loe6, and lo-’* 

0 The corresponding recommended criteria are 27 ug/L, 2.7 ug/L, and I_ 

0.27 ug/L, respectively. If these estimates are made for 
consumption of aquatic organisms only, excluding consumption 

of water, the levels are 807 ug/L, 80.7 ug/L, and 8.07 ug/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERIA: 

VINYL CHLORIDE 

Aquatic Life 

No freshwater organisms have been tested with vinyl chloride 

and no statement can be made concerning acute or chronic toxicity. 

No saltwater organisms have been tested with vinyl 

chloride and no statement can be made concerning acute or 

chronic toxicity. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to vinyl chloride through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, based 

on the nonthreshold assumption for this chemical. However, 

zero level may not be attainable at the present time. 

Therefore, the levels which may result in incremental increase of 

cancer risk over the lifetime are estimated at 10-5r and 

The corresponding recommended criteria are 20 ug/L, 2.0 

ug/L, and 0.2 ug/L, respectively. If these estimates are made 

for consumption of aquatic organisms only, excluding consumption 

of water, the levels are 5,246 ug/L, 525 ug/L, and 52.5 ug/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
ZINC 

CRITERIA: 

Aquatic - Life 

For total recoverable zinc the crit rion to protect fre hwater 

aquatic life as derived using the Guidelines is 47 ug/L as a 24- 

hour average and the concentration (in ug/L) should not 

exceed the numerica 1 value given by 

0.83 [ ln(hardness) ]+1.95) at any time. For example, at e( 

hardnesses of 50, 100, and 200 mg/L as CaC03 the 

concentration of total recoverable zinc should not exceed 180, 

320, and 570 ug/L at any time. 

For total recoverable zinc the criterion to protect saltwater 

aquatic life as derived using the Guidelines is 58 ug/L as a 24- 

hour average and the concentration should not exceed 190 ug/L at 

any time. 

Human Health 

Sufficient data are not available for zinc to derive a level 

which would protect against the potential toxicity of this 

compound. Using available organoleptic data, to control 

undesirable taste and odor quality of ambient water the estimated 

level is 5 mg/L. It should be recognized that organoleptic data 

have limitations as a basis for establishing a water quality 

criteria, and have no demonstrated relationship to potential 

adverse human health effects. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
I ~ 

APPENDIX A 
DERIVATION -

OF -THE - 1985 CRITERION 

Derivation of numerical national water quality criteria for 

the protection of aquatic organisms and their uses is a complex 

process that uses information from many areas of aquatic 

toxicology. After a decision is made that a national criterion 

is needed for a particular material, all available information 

concerning toxicity to, and bioaccumulation by, aquatic organisms 

is collected, reviewed for acceptability, and sorted. If enough 

acceptable data on acute toxicity to aquatic animals are 

available, they are used to estimate the highest 1-hour average 

concentration that should not result in unacceptable effects on 

aquatic organisms and their uses. If justified, this 

concentration is made a function of a water quality 

characteristic such as pH, salinity, or hardness. Similarly, 

data on the chronic toxicity of the material to aquatic animals 

are used to estimate the highest 4-day average concentration 

that should not cause unacceptable toxicity during a long-term 

exposure. If appropriate, this concentration is also related to 

a water quality characteristic. 

Data on toxicity to aquatic plants are examined to determine 

whether plants are likely to be unacceptably affected by 

concentrations that should not cause unacceptable effects on 

animals. Data on bioaccumulation by aquatic organisms are used 

to determine if residues might subject edible species to 

restrictions by the U.S. Food and Drug Administration or if such 

residues might harm some wildlife consumers of aquatic life. All 

other available data are examined €or adverse effects that might 
be biologically important. 

If a thorough review of the pertinent information indicates 

that enough acceptable data are available, numerical national 

water quality criteria are derived for fresh water or saltwater 

or both to protect aquatic organisms and their uses from 

unacceptable effects due to exposures to high concentrations for 

short periods of time, lower concentrations for longer periods of 

time, and combinations of the two. 

- I. Collection -- of Data 

A. Collect all available data on the material 

concerning (a) toxicity to, and biqaccumulation 

by, aquatic animals and plants, (b) FDA action 

levels [12], and (c) chronic feeding studies and 

long-term field studies with wildlife species that 

regularly consume aquatic organisms. 

B. All data that are used should be available in 

typed, dated, and signed hard copy (pub1 ication, 

manuscript, letter, memorandum, etc.) with enough 

supporting information to indicate that acceptable 

test procedures were used and that the results are 

probably reliable. In some cases it may be 

appropriate to obtain additional written 

information from the investigator, if possible. 

Information that is confidential or privileged or 

otherwise not available for distribution should 

not be used. 

C. Questionable data, whether published or 
unpublished, should not be used. For example, data 

should usually be rejected if they are from tests 

that did not contain a control treatment, tests in 

which too many organisms in the control treatment 

died or showed signs of stress or disease, and 

tests in which distilled or deionized water was 

used as the dilution water without addition of 

appropriate salts. 

D. Data on technical grade materials may be used if 

appropriate, but data on formulated mixtures and 

emulsifiable concentrates of the material of 

concern should not be used. 

E. For some highly volatile, hydrolyzable, or 

degradable materials it is probably appropriate to 

use only results of flow-through tests in which 

the concentrations of test material in the test 

solutions were measured often enough using 

acceptable analytical methods. 

F. Data should be rejected if they were obtained 

using: 

1. Brine shrimp, because they usually occur 

naturally only in water with salinity greater 

than 35 g/kg. 

2. Species that do not have reproducing wild 

populations in North America (See Appendix 1). 

3. Organisms that were previously exposed to 0 t. substantial concentrations of the test 

material or other contaminants. 
G. Questionable data, data on formulated mixtures and 

emulsifiable concentrates, and data obtained with 

nonresident species or previously exposed 

organisms may be used to provide auxiliary 

information but should not be used in the 

derivation of criteria. 

- 11. Required - Data 

A. Certain data should be available to help ensure 

that each of the four major kinds of possible 

adverse effects receives adequate consideration. 

Results of acute and chronic toxicity tests with 

representative species of aquatic animals are 

necessary so that data available for tested 

species can be considered a useful indication of 

the sensitivities of appropriate untested species. 

Fewer data concerning toxicity to aquatic plants 

are required because procedures for conducting 

tests with plants and interpreting the results of 

such tests are not as well developed. Data 

concerning bioaccumulation by aquatic organisms 

are required only if relevant data are available 

concerning the significance of residues in aquatic 

organisms. 

I 

B. To derive a criterion for freshwater aquatic 

organisms and their uses, the following should be 

available: 

1. Results of acceptable acute tests (see Section 
IV) with at least one species of freshwater 

animal in at least eight different families 

such that all of the following are included: 

a. 

b. 

C. 

d. 

e. 

f. 

g. 

h. 

the family Salmonidae in the class 

Osteichthyes 

a second family in the class 

Osteichthyes, preferably a commercially or 

recreationally important warmwater species 

(e.g., bluegill, channel catfish, etc.) 

a third family in the phylum Chordata (may 

be in the class Osteichthyes or may be an 

amphibian, etc.) 

a planktonic crustacean (e.g., cladoceran, 

copepod, etc.) 

a benthic crustacean (e.g., ostracod, 

isopod, amphipod, crayfish, etc.) 

an insect (e.g., mayfly, dragonfly, 

damselfly, stonefly, caddis fly, mosquito, 

midge, etc.) 

a family in a phylum other than Arthropoda 

or Chordata (e.g., Rotifera, Annelida, 

Mol lusca, etc.) 

a family in any order of insect or any 

phylum not already represented. 

2. Acute-chronic ratios (see Section TI) with 

species of aquatic animals in at least three 

different families provided that of the three 

species: 
a. at least one is a fish 

b. at least one is an invertebrate 

c. at least one is an acutely sensitive 

freshwater species (the other two may be 

saltwater species). 

3. Results of at least one acceptable test with a 

freshwater alga or vascular plant (see Section 

VIII). If plants are among the aquatic 

organisms that are most sensitive to the 

material, results of a test with a plant in 

another phylum (division) should also be 

available. 

4. At least one acceptable bioconcentration 

factor dete'rmined with an appropriate 

freshwater species, if a maximum permissible 

tissue concentration is available (see Section 

IX) . C. To derive a criterion for saltwater aquatic 

organisms and their uses, the following should be 

available: 

1. Results of acceptable acute tests (see Section 

IV) with at least one species of saltwater 

animal in at least eight different families 

such that all of the following are included: 

a. two families in the phylum Chordata 

b. a family in a phylum other than Arthropoda 

or Chordata 
c. either the Mysidae or Penaeidae family 

d. three other families not in the phylum 

Chordata (may include Mysidae or 

Penaeidae, whichever was not used above) 

e. any other family. 

2. Acute-chronic ratios (see section VI) with 

species of aquatic animals in at least three 

different families provided that of the three 

species: 

a. at least one is a fish 

b. at least one is an invertebrate 

c. at least one is an acutely sensitive 

saltwater species (the other one may be a 

freshwater species). 

3. Results of at least one acceptable test with a 

saltwater alga or vascular plant (see Section 

VIII. If plants are among the aquatic 

organisms most sensitive to the material, 

results of a test with a plant in another 

phylum (division) should also be available. 

4. At least one acceptable bioconcentration 

factor determined with an appropriate 

saltwater species, if a maximum permissible 

tissue concentration is available (see Section 

IX) * 

D. If all the required data are available, a numerical 0 criterion can usually be derived, except in special %. 

cases. For example, derivation of a criterion 
might not be possible if the available acutechronic ratios vary by more than a factor of 10 

with no apparent pattern. Also, if a criterion is 

to be related to a water quality characteristic T 

(see Sections V and VII) , more data will be 

necessary. 

Similarly, if all required data are not available, 

a numerical criterion should not be derived except 

in special cases. For example, even if not enough 

acute and chronic data are available, it might be 

possible to derive a criterion if the available 

data clearly indicate that the Final Residue Value 

should be much lower than either the Final Chronic 

Value or the Final Plant Value. 

E. Confidence in a criterion usually increases as the 

amount of available pertinent data increases. 

Thus, additional data are usually desirable. 

111. Final Acute Value 

A. Appropriate measures of the acute (short-term) 

toxicityofthematerialtoavarietyofspecies of 

aquatic animals are used to calculate the Final 

Acute Value. The Final Acute Value is an estimate 

' of the concentration of the material corresponding 

to a cumulative probability of 0.05 in the acute 

toxicity values for the genera with which 

acceptable acute tests have been conducted on the 

material. However, in some cases, if the Species 
Mean Acute Value of a commercially or 

recreational 1 y important species is lower than the 

calculated Final Acute Value, then that Species 

Mean Acute Value replaces the calculated Final 

Acute Value in order to provide protection for that 

important species. 

B. Acute toxicity tests should have been conducted 

using acceptable procedures [13]. 

C. Except for tests with saltwater annelids and 

mysids, results of acute tests during which the 

test organisms were fed should not be used, unless 

data indicate that the food did not affect the 

toxicity of the test material. 

D. Results of acute tests conducted in unusual 0 

dilution water, e.g., dilution water in which total 

organic carbon or particulate matter exceeded 5 

mg/L, should not be used, unless a relationship is 

developed between acute toxicity and organic carbon 

or particulate matter or unless data show that 

organic carbon, particulate matter, etc., do not 

affect toxicity. 

E. Acute values should be based on endpoints which 

reflect the total severe acute adverse impact of 

the test material on the organisms used in the 

test. Therefore, only the following kinds of data 

on acute toxicity to aquatic animals should be 

used: 
1. Tests with daphnids and other cladocerans 

should be started with organisms less than 24 

hours old and tests with midges should be 

stressed with second- or third -instar larvae. 

The result should be the 48-hr EC50 based on 

percentage of organisms immobilized plus 

percentage of organisms killed. If such an 

EC50 is not available from a test, the 48-hr 

LC50 should be used in place of the desired 

48-hr EC50. An EC50 or LC50 of longer than 

48 hours can be used as long as the animals 

were not fed and the control animals were 

acceptable at the end of the test. 

a 2. The result of a test with embryos and larvae of barnacles, bivalve molluscs (clams, 

mussels, oysters, and scallops) , sea urchins, 

lobsters, crabs, shrimp, and abalones should 

be the 96-hr EC50 based on the percentage of 

organisms with incompletely developed shells 

plus the percentage of organisms killed. If 

such an EC50 is not available from a test, the 

lower of the 9 6 - hr EC50 based on the 

percentage of organisms with incompletely 

developed shells and the 96-hr LC50 should be 

used in place of the desired 96-hr EC50. If 

the duration of the test was between 48 and 96 

hours, the EC50 or LC50 at the end of the test 

should be used. 
3. The acute values from tests with all other 

freshwater and saltwater animal species and 

older life stages of barnacles, bivalve 

molluscs, sea urchins, lobsters, crabs, 

shrimps, and abalones should be the 96-hr EC50 

based on the percentage of organisms 

exhibiting loss of equilibrium plus the 

percentage of organisms immobilized plus the 

percentage OF organisms killed. If such an 

EC50 is not available from a test, the 96-hr 

LC50 should be used in place of the desired 

96-hr EC50. 

4. Tests with single - celled organisms are not 

considered acute testa, even if the duration 

was 96 hours or less. 

0 

5. If the tests were conducted properly, acute 

values reported as " greater than" values and 

those which are above the solubility of the 

test material should be used, because 

rejection of such acute values would 

unnecessarily lower the Final Acute Value by 

eliminating acute values for resistant 

species. 

F. If the acute toxicity of the material to aquatic 

animals apparently has been shown to be related to 

a water quality characteristic such as hardness or 

particulate matter for freshwater animals or 
salinity or particulate matter for saltwater 

animals, a Final Acute Equation should be derived 

based on that water quality characteristic. Go to 

Section V. 

G. If the available data indicate that one or more 

life stages are at least a factor of 2 more 

resistant than one ormore other life stages ofthe 

same species, the data for the more resistant life 

stages should not be used in the calculation of the 

Species Mean Acute Value (SMAV) because a species 

can only be considered protected from acute 

toxicity if all life stages are protected. 

H. The agreement of the data within and between 

species should be considered. Acute values that 

appear to be questionable in comparison with other 

acute and chronic data for the same species and for 

other species in the same genus probably should not 

be used in calculation of a Species Mean Acute 

Value. For example, if the acute values available 

For a species or genus differ by more than a factor 

of 10, some or all of the values probably should 

not be used in calculations. 

I. For each species for which at least one acute 

value is available, the Species Mean Acute Value 

should be calculated as the geometric mean of the 

results of all flow-through tests in which the 

concentrations of test material were measured. For 

a species for which no such result is available, 
the Species Mean Acute Value should be calculated 

as the geometric mean of all available acute 

values, i.e., results of flow-through tests in 

which the concentrations were not measured and 

results of static and renewal tests based on 

initial concentrations of test material (nominal 

concentrations are acceptable for most test 

materials if measured concentrations are not 

available). 

NOTE: Data reported by original investigators should not 

be rounded off. Results of all intermediate 

calculations should be rounded [14] to four 

significant digits. 

0 NOTE: The geometric mean of N numbers is the Nth root of 

the product of the N numbers. Alternatively, the 

geometric mean can be calculated by adding the 

logarithms of the N numbers, dividing the sum by 

N, and taking the antilog of the quotient. The 

geometric mean of two numbers is the square root 

of the product of the two numbers, and the 

geometric mean of one number is that number. 

Either natural (base 0) or common (base 10) 

logarithms can be used to calculate geometric 

means as long as they are used consistently within 

each set of data, i.e., the antilog used must 

match the logarithm Used. 
NOTE: Geometric means, rather than arithmetic means, are 

used here because the distributions of 

sensitivities of individual organisms in toxicity 

tests on most materials and the distributions of 

sensitivities of species within a genus are more 

likely to be lognormal than normal. Similarly, 

geometric means are used for acute-chronic ratios 

and bioconcentration factors because quotients are 

likely to be closer to lognormal than normal 

distributions. In addition, division of the 

geometric mean of a set of numerators by the 

geometric mean of the set of corresponding 

denominators will result in the geometric mean of 

the set of corresponding quotients. 

J. For each genus €or which one or more Species Mean 

Acute Values are available, the Genus Mean Acute 

Value should be calculated as the geometric mean 

of the Species Mean Acute Values available for the 

genus. 

K. Order the Genus Mean Acute Value from high to low. 

L. Assign ranks, R, to the Genus Mean Acute Value 

from vvlvv for the lowest to *'N" €or the highest. 

If two or more Genus Mean Acute Values are 

identical, arbitrarily assign them successive 

ranks. 

M. Calculate the cumulative probability, P, €or each 

Genus Mean Acute Value as R/ (N+l). 
N. Select the four Genus Mean Acute Value which have 

cumulative probabilities closest to 0.05 (if there 

are less than 59 Genus Mean Acute Value, these 

will always be the four lowest Genus Mean Acute 

Values). 

Using the selected Genus Mean Acute Values and Fs, 

calculate: 

S2= E(ln GMAV)2)- ((Eln GMAV))2/4) 

(PI - ((E /"~))2/4) 

0. 

L = (E(1n GMAV) - S(E(/Ap)))/4 

A = S(/"O.OS) +L 

FAV = eA 

(See [113 for development of the calculation 

procedure and Appendix 2 for example calculation 

and computer program.) 

NOTE: Natural logarithms (logarithms to base el denoted 

as In) are used herein merely because they are 

easier to use on some hand calculators and 

computers than common (base 10) logarithms. 

Consistent use of either will produce the same 

result. 

P. If for a commercially or recreationally important 

species the geometric mean of the acute values 

from flow - through tests in which the 

concentrations of test material were measured is 

lower than the calculated Final Acute Value, then 

that geometric mean should be used as the Final 

Acute Value instead of the calculated Final Acute 
Value. 

Q. GO to section VI. 
- IV. Final Acute Equation 0 

A. When enough data are available to show that acute 

toxicity to two or more species is similarly 

related to a water quality characteristic, the 

relationship should be taken into account as 

described in Sections B-G below or using analysis 

of covariance [15,16]. The two methods are 

equivalent and produce identical results. The 

manual method described below provides an 

unuerstanding of this application of covariance 

analysis, but computerized versions of covariance 

analysis are much more convenient for analyzing 

large data tests. If two or more factors affect 

toxicity, multiple regression analysis should be 

used. 

B. For each species for which comparable acute 

toxicity values are available at two or more 

different values of the water quality 

characteristic, perform a least squares regression 

of the acute toxicity values on the corresponding 

values of the water quality characteristic to 

obtain the slope and its 95 percent confidence 

limits €or each species. 

NOTE: Because the best documented relationship fitting 

these data is that between hardness and acute 

toxicity of metals in fresh water and a log-log 
relationship, geometric means and natural 

logarithms of both toxicity and water quality are 

used in the rest of this section. For 

relationships based on other water quality 

characteristics such as pH, temperature, or 

salinity, no transformation or a different 

transformation might fit the data better, and 

appropriate changes will be necessary throughout 

this section. 

C. Decide whether the data for eachspecies are 

useful, taking into account the range and number 

of the tested values of the water quality 

characteristic and the degree of agreement within 

and between species. For example, a slope based 

on six data points might be of limited value if it 

is based only on data for a very narrow range of 

values of the water quality characteristic. A 

slope based on only two data points, however, 

might be useful if it is consistent with other 

information and if the two points cover a broad 

enough range of the water quality characteristic. 

In addition, acute values that appear to be 

questionable in comparison with other acute and 

chronic data available for the same species and 

for other species in the same genus probably 

should not be used. For example, if after 

adjustment for the water quality characteristic, 

the acute values available for a species or genus 
differ by more than a factor of 10, probably some 

or all of the values should be rejected. If 

useful slopes are not available for at least one 

fish and one invertebrate or if the available 

slopes are too dissimilar or if too few data are 

available to adequately define the relationship 

between acute toxicity and the water quality 

characteristic, return to Section IV.G, using the 

results of tests conducted under conditions and in 

waters similar to those commonly used for toxicity 

tests with the species. 

D. Individually for each species calculate the 

geometric mean of the available acute values and 

then divide each of the acute values for species 

by the mean for the species. This normalizes the 

values so that the geometric mean of the 

normalized values for each species individual1 y 

and for any combination of species is 1.0. 

E. Similarly normalize the values of the water 

quality characteristic for each species 

individually. 

F. Individually for each species perform a least 

squares regression of the normalized acute 

toxicity values on the corresponding normalized 

values of the water quality characteristic. The 

resulting slopes and 95 percent confidence limits 

~. will be identical to those obtained in Section B. 
NOW, however, if the data are actually plotted, 

the line of best fit for each individual 

species will go through the point 1,l in the 

center of the graph. 

G. Treat all the normalized data as if they were all 

for the same species and perform a least squares 

regression of all the normalized acute values on 

the corresponding normalized values of the water 

quality characteristic to obtain the pooled acute 

slope, V, and its 95 percent confidence limits. 

If all the normalized data are actually plotted, 

the line of best fit will go through the point 1,l 

in the center of the graph. 

H. For each species aalculate the geometric mean, W, 

of the acute toxicity values and the geometric 

mean, XI of the values of the water quality 

characteristic. (These were calculated in steps D 

and E.) 

I. For each species calculate the logarithm, Y, of 

the Species Mean Acute Value at a selected value, 

2, of the water q u a l i t y c h a r a c t e r i s t i c using the 

equation: 

Y = In W - v(ln X - In 2). 

J. For each species calculate the SMAV at Z using 

the equation: SMAV = eY. 

NOTE: Alternatively, the Species Mean Acute Values at Z 

can be obtained by skipping step H using the 
equations in steps I and J to adjust each acute 

value individually to 2, and then calculating the 

geometric mean of the adjusted values €or each 

species individually. This alternative procedure 

allows an examination of the range of the adjusted 

acute values for each species. 

K. Obtain the Final Acute Value at Z by using the 

procedure described in Section 1V.J-0. 

L. If the Species Mean Acute Value at Z of a 

commercially or recreationally important species 

is lower than the calculated Final Acute Value at 

Z, then that Species Mean Acute Value should be 

used as the Final Acute Value at Z instead of the 

calculated Final Acute Value. 

M. The Final Acute Equation is written as: Final 

Acute Value = .(V[ln(water quality 

characteristic)] + In A - V[ln Z]), where V = 

pooled acute slope andA=Final Acutevalueat 2. 

Because V, A, and 2 are known, the Final Acute 

Value can be calculated for any selected value of 

the water quality characteristic. 

V. - Final Chronic Value 

A. Depending on the data that are available 

concerning chronic toxicity to aquatic animals, 

the Final Chronic Value might be calculated in the 

same manner as the Final Acute Value or by 

dividing the Final Acute Value by the Final Acute- 
NOTE : 

B. 

Chronic Ratio. In some cases it may not be 

possible to calculate a Final Chronic Value. 

As the name implies, the acute-chronic ratio is a 

way of relating acute and chronic toxicities. The 

acute - chronic ratio is basically the inverse of 

the application factor, but this new name is 

better because it is more descriptive and should 

help prevent confusion between 'application 

factors" and "safety factors." Acute- chronic 

ratios and application factors are ways of 

relating the acute and chronic toxicities of a 

material to aquatic organisms. Safety factors are 

used to provide an extra margin of safety beyond 

the known or estimated sensitivities of aquatic 

organisms. Another advantage of the acute-chronic 

ratio is that it will usually be greater than 1; 

this should avoid the confusion as to whether a 

large application factor is one that is close to 

unity or one that has a denominator that is much 

greater than the numerator. 

Chronic values should be based on results of flowthrough (except renewal is acceptable for 

daphnids) chronic tests in which the 

concentrations of test material in the test 

solutions were properly measured at appropriate 

I 

times during the test. 

.* C. Results of chronic tests in which survival, 
growth, or reproduction in the control treatment 

was unacceptably low should not be used. The 

limits of acceptability will depend on the 

species. 

D. Results of chronic tests conducted in unusual 

dilution water, e.g., dilution water in which 

total organic carbon or particulate matter 

exceeded 5 mg/L, should not be used, unless a 

relationship is developed between chronic toxicity 

and organic carbon or particulate matter or 

unless data show that organic carbon, particulate 

matter, etc., do not affect toxicity. 

E. Chronic values should be based on endpoints and 

lengths of exposure appropriate to the species. 

Therefore, only results of the following kinds of 

chronic toxicity tests should be used: 

1. Life - cycle toxicity tests consisting of 

exposures of each of two or more groups of 

individuals of a species to a different 

concentration of the test material throughout 

a life cycle. To ensure that all life stages 

and life processes are exposed, tests with 

fish should begin with embryos or newly 

hatched young less than 48 hours old, continue 

through maturation and reproduction, and 

should end not less than 24 days (90 days for 

salmonids) after the hatching of the next 

generation. Tests with daphnids should begin 
with young less than24 hours oldand last for 

not less than 21 days. Tests with mysids 

should begin with young less than 24 hours old 

and continueuntil 7 days past themediantime 

of first brood release in the controls. For 

fish, data should be obtained and analyzed on 

survival and growth of adults and young, 

maturation of males and females, eggs spawned 

per female, embryo viability (salmonids only) , 

and hatchability. For daphnids, data should 

be obtained and analyzed on survival and young 

per female. For mysids, data should be 

obtained and analyzed on survival, growth, and 

young per female. 

2. Partial life - cycle toxicity tests consisting 

of exposures of each of two or more groups of 

individuals of a species of fish to a 

concentration of the test material through 

most portions of a life cycle. Partial lifecycle tests are allowed with fish species that 

require more than a year to reach sexual 

maturity, so that all major life stages can be 

exposed to the test material in less than 15 

months. Exposure to the test material should 

begin with immature juveniles at least 2 

months prior to active gonad development, 

continue through maturation and reproduction, 
and end not less than 24 days (90 days for 

salmonids) after the hatching of the next 

generation. Data should be obtained and 

analyzed on survival and growth of adults and 

young, maturation of males and females, eggs 

spawned per female, embryo viability 

(salmonids only), and hatchability. 

3. Early life-stage toxicity tests consisting of 

28- to 32-day (60 days. post hatch for 

salmonids) exposures of the early life stages 

of a species of fish from shortly a,fter 

fertilization through embryonic, larval, and 

early juvenile development. Data should be 

obtained and analyzed on survival and growth. 

NOTE: Results of an early life-stage test are used as 

predictions of results of life-cycle and partial 

life-cycle tests with the same species. 

Therefore, when results of a life-cycle or partial 

life-cycle test are available, results of an early 

life-stage test with the same species should not 

be used. Also, results of early life-stage tests 

in which the incidence of mortalities or 

abnormalities increased substantially near the end 

of the test should not be used because results of 

such tests are possibly not good predictions of 

the results of comparable life-cycle or partial 

1 if e-cycle tests. 
0 F. A chronic value may be obtained by calculating the 

geometric mean of the lower and upper chronic 

limits from a chronic test or by analyzing chronic 

data using regression analysis. A lower chronic 

limit is the highest tested concentration (a) in 

an acceptable chronic test, (b) which did not 

cause an unacceptable amount of adverse effect on 

any of the specified biological measurements, and 

(c) below which no tested concentration caused an 

unacceptable effect. An upper chronic limit is 

the lowest tested concentration (a) in an 

acceptable chronic test, (b) which did cause an 

unacceptable amount of adverse effect on one or 

more of the specified biological measurements, and 

(c) above which all tested concentrations also 

caused such an effect. 

NOTE: Because various authors have used a variety of 

terms and definitions to interpret and report 

results of chronic tests, reported results should 

be reviewed carefully. The amount of effect that 

is considered unacceptable is often based on a 

statistical hypothesis test, but might also be 

defined in terms of a specified percent reduction 

from the controls. A small percent reduction 

(e.g., 3 percent) might be considered acceptable 

even if it is statistically significantly 

different from the control, whereas a large 
percent reduction (e.g., 30 percent) might be 

considered unacceptable even if it is not 

statistically significant. 

G. If the chronic toxicity of the material to aquatic 

animals apparently has been shown to be related to 

a water quality characteristic such as hardness or 

particulate matter for freshwater animals or 

salinity or particulate matter for saltwater 

animals, a Final Chronic Equation should be 

derived based on that water quality 

characteristic. Go to Section VII. 

H. If chronic values are available for species in 

eight families as described in Sections III.B.1 or 

III.C.1, a Species Mean Chronic Value (SMCV) 

should be calculated for each species for which at 

least one chronic value is available by 

calculating the geometric mean of all chronic 

values available for the species, and appropriate 

Genus Mean Chronic Values should be calculated. 

The Final Chronic Value should then be obtained 

using the procedure described in Section 1V.J- 0. 

Then go to Section V1.M. 

I. For each chronic value for which at least one 

corresponding appropriate acute value is 

available, calculate an acute-chronic ratio, using 

for the numerator the geometric mean of the 0 results of all acceptable f low-through (except 

static is acceptable for daphnids) acute tests in 
the same dilution water and in which the 

concentrations were measured. For fish, the acute 

test(s) should have been conducted with juveniles. 

The acute test(s) should have been part of the 

same study as the chronic test. If acute tests 

were not conducted as part of the same study, 

acute tests conducted in the same laboratory and 

dilution water, but in a different study, may be 

used. If no such acute tests are available, 

results of acute tests conducted in the same 

dilution water in a different laboratory may be 

used. If no such acute tests are available, an 

acute-chronic ratio should not be calculated. 

J. For each species, calculate the species mean 

acute-chronic ratio as the geometric mean of all 

acute-chronic ratios available for that species. 

K. For some materials the acute-chronic ratio seems 

to be the same for all species, but for other 

materials the ratio seems to increase or decrease 

as the Species Mean Acute Value (SMAV) increases. 

Thus the Final Acute-Chronic Ratio can be obtained 

in four ways, depending on the data available: 

1. If the Species Mean Acute-Chronic ratio Seems 

to increase or decrease as the Species Mean 

Acute Value increases, the Final Acute-Chronic 

Ratio should be calculated as the geometric 

mean of the acute-chronic ratios for species 
whose Species Mean Acute Values are close to 

the Final Acute Value. 

2. If no major trend is apparent and the acutechronic ratios for a number of species are 

within a factor of 10, the Final AcuteChronic Ratio should be calculated as the 

geometric mean of all the Species Mean AcuteChronic Ratios available for both freshwater 

and saltwater species. 

3. For acute tests conducted on metals and 

possibly other substances with embryos and 

larvae of barnacles, bivalve molluscs, sea 

urchins, lobsters, crabs, shrimp, and abalones 

(see Section IV.E.2), it is probably 0 

appropriate to assume that the acute - chronic 

ratio is 2. Chronic tests are very difficult 

to conduct with most such species, but it is 

likely that the sensitivities of embryos and 

larvae would determine the results of lifecycle tests. Thus, if the lowest available 

Species Mean Acute Values were determined with 

embryos and larvae of such species, the Final 

Acute-Chronic Ratio should probably be assumed 

to be 2, so that the Final Chronic Value is 

equal to the Criterion Maximum Concentration 

(see Section X1.B) a *.,~ ..?a. ...,.. 
4. If the most appropriate Species Mean AcuteChronic Ratios are less than 2.0, and 

especially if they are less than 1.0, 

acclimation has probably occurred during the 

chronic test. Because continuous exposure and 

acclimation cannot be assured to provide 

adequate protection in field situations, the 

Final Acute-Chronic Ratio should be assumed to 

be 2, so that the Final Chronic Value is equal 

to the Criterion Maximum Concentration (see 

Section X1.B). 

If the available Species Mean Acute-Chronic 

Ratios do not fit one of these cases, a Final 

Acute-Chronic Ratio probably cannot be 

obtained, and a Final Chronic Value probably 

cannot be calculated. 

L. Calculate the Final Chronic Value by dividing the 

Final Acute Value by the Final Acute-Chronic 

Ratio. If there was a Final Acute Equation rather 

than a Final Acute Value, see also Section VI1.A. 

M. If the Species Mean Chronic Value of a 

commercially or recreational ly important species 

is lower than the calculated Final Chronic Value, 

then that species Mean Chronic Value should be 

used as the Final Chronic Value instead of the 

calculated Final Chronic Value. 

N. Go to Section VIII. 
VI. Final Chronic Equation 0- A. A Final Chronic Equation can be derived in two 

ways. The procedure described here in Section A 

will result in the chronic slope being the same as 

the acute slope. The procedure described in 

Sections B-N usually will result in the chronic 

slope being different from the acute slope. 

1. If acute-chronic ratios are available €or 

enough species at enough values of the water 

quality characteristic to indicate that the 

acute-chronic ratio is probably the same for 

all species and is probably independent of the 

water quality characteristic, calculate the 

Final Acute-Chronic Ratio as the geometric 

mean of the available Species Mean AcuteChronic Ratios. 

’ 

2. Calculate the Final Chronic Value at the 

selected value Z of the water quality 

characteristic by dividing the Final Acute 

Value at Z (see Section V.M) by the Final 

Acute-Chronic Ratio. 

3. Use V = pooled acute slope (see section V.M) 

as L = pooled chronic slope. 

4. Go to Section VI1.M. 

B. When enough data are available to show that 

chronic toxicity to at least one species is 

related to a water quality characteristic, the 
relationship should be taken into account as 

described in Sections B-G or using analysis of 

covariance [15,16]. The two methods are 

equivalent and produce identical results. The 

manual method described below provides an 

understanding of this application of covariance 

analysis, but computerized versions of covariance 

analysis are much more convenient for analyzing 

large data sets. If two or more factors affect 

toxicity, multiple regression analysis should be 

used. 

For each species for which comparable chronic 

toxicity values are available at two or more 

different values of the water quality 

characteristic, perform a least squares regression 

of the chronic toxicity values on the 

corresponding values of the water quality 

characteristic to obtain the slope and its 95 

percent confidence limits for each species. 

NOTE: Because the best documented relationship fitting 

these data is that between hardness and acute 

toxicity of metals in freshwater and a log-log 

relationship, geometric means and natural 

logarithms of both toxicity and water quality are 

used in the rest of this section. For 

relationships based on other water quality 

characteristics such as pH, temperature, or 
salinity, no transformation or a different 

transformation might fit the data better, and 

appropriate changes will be necessary throughout 

this section. ft is probably preferable, but not 

necessary, to use the same transformation that was 

usedwiththeacutevalues insectionv. 

D. Decide whether the data for each speciesare 

useful, taking into account the range and number 

of the tested values of the water quality 

characteristic and the degree of agreement within 

and between species. For example, a slope based 

on six datapointsmightbeof limitedvalue if it 

is based only on data €or a very narrow range of 

values of the water quality characteristic, A 

slope based on only two data points, however, 

might be useful if it is consistent with other 

information and if the two points cover a broad 

enough range of the water quality characteristic. 

In addition, chronic values that appear to be 

questionable in comparison with other acute and 

chronic data available €or the same species and 

for other species in the same genus probably 

should not be used. For example, if after 

adjustment for the water qua1 ity characteristic, 

the chronic values available for a species or 

genus differ by more than a factor of 10, probably 

some or all of the values should be rejected. If 
a useful chronic slope is not available for at 

least one species or if the available slopes are 

too dissimilar or if too few data are available to 

adequately define the relationship between chronic 

toxicity and the water quality characteristic, it 

might be appropriate to assume that the chronic 

slope is the same as the acute slope, which is 

equivalent to assuming that the acute-chronic 

ratio is independent of the water quality 

characteristic. Alternatively, return to Section 

VI.H, using the results of tests conducted under 

conditions and in waters similar to those commonly 

used for toxicity tests with the species. 

E. Individually for each species calculate the 

geometric mean of the available chronic values and 

then divide each chronic value for a species by 

the mean for the species. This normalizes the 

chronic values so that the geometric mean of the 

normalized values for each species individually 

and for any combination of species is 1.0. 

F. Similarly normalize the values of the water 

quality characteristic for each species 

individually. 

G. Individually for each species perform a least 

squares regression of the normalized chronic 

toxicity values on the corresponding normalized 

values of the water quality characteristic. The 

resulting slopes and the 95 percent confidence 
limits will be identical to those obtained in 

Section B. Now, however, if the data are actually 

plotted, the line of best fit for each individual 

species will go through the point 1,1 in the 

center of the graph. 

K. Treat all the normalized data as if they were all 

for the same species and perform a least squares 

regression of all the normalized chronic values on 

the corresponding normalized values of the water 

quality characteristic to obtain the pooled 

chronic slope, L, and its 95 percent confidence 

limits, If all the normalized data are actually 

plotted, the line of best fit will go through the 

point 1,l in the center of the graph. 0 I. For each species calculate the geometric mean, MI 

of the toxicity values and the geometric mean, PI 

of the values of the water quality characteristic. 

(These were calculated in steps E and F.) 
J. 

NOTE : 

K. 

NOTE : 

L. 

M. 

For each species calculate the logarithm, Q, of 

the Species Mean Chronic Value at a selected 

value, Z, of the water quality characteristic 

using the equation: Q = In M - L(ln P - In Z). 

Although it is not necessary, it will usually be 

best to use the same value of the water quality 

characteristic here as was used in section V.I. 

For each species calculate a Species Mean Chronic 

Value at z using the equation: SMCV = eQ. 

Alternatively, the Species Mean Chronic Values at 

Z can be obtained by skipping step J, using the 

equations in steps J and K to adjust each acute 

value individually to 2, and then calculating the 

geometric means of the adjusted values for each 

species individually. This alternative procedure 

allows an examination of the range of the adjusted 

chronic values for each species. 

Obtain the Final Chronic Value at 2 by using the 

procedure described in Section 1V.J-0. 

If the Species Mean Chronic Value at Z of a 

commercially or recreationally important species 

is lower than the calculated Final Chronic Value 

at Z, then that Species Mean Chronic Value should 

beused as the Final ChronicValueatZ insteadof 

the calculated Final Chronic Value. 
N. The Final Chronic Equation iS written as: Final 

Chronic Value = e (L[ln(water quality 

characteristic)] + In S - L[ln Z]), where L = 

pooled chronic slope and S = Final Chronic Value 

at 2. Because L, S and Z are known, the Final 

Chronic Value can be calculated for any selected 

value of the water quality characteristic. 

Final Plant Value 

A. Appropriate measures of the toxicity of the 

material to aquatic plants are used to compare the 

relative sensitivities of aquatic plants and 

animals. Although procedures for conducting and 

interpreting the results of toxicity tests with 

plants are not well developed, results of tests 

with plants usually indicate that criteria which 

adequately protect aquatic animals and their uses 

will probably also protect aquatic plants and 

their uses. 

B. A plant value is the result of a 96-hr test 

conducted with an alga or a chronic test conducted 

with an aquatic vascular plant. 

NOTE: A test of the toxicity of a metal to a plant 

usually should not be used if the medium contained 

an excessive amount of a complexing agent, such as 

EDTA, that might affect the toxicity of the metal. 

Concentrations of EDTA above about 200 ug/L should 

probably be considered excessive. 
C. The Final Plant Value should be obtained by 

selecting the lowest result from a test with an 

important aquatic plant species in which the 

concentrations of test material were measured and 

the endpoint was biologically important. 

VIII. Final Residue Value 

A. The Final Residue Value is intended to (a) prevent 

concentrations in commercially or recreationally 

important aquatic species from affecting 

marketability because of exceedence of applicable 

FDA action levels and (b) protect wildlife, 

including fishes and birds, that consume aquatic 

organisms from demonstrated unacceptable effects. 

The Final Residue Value is the lowest of the 

residue values that are obtained by dividing 

maximum permissible tissue concentrations by 

appropriate bioconcantration or bioaccumulation 

factors. A maximum permissible tissue 

concentration is either (a) an FDA action level 

[12] for fish oil or for the edible portion of 

fish or shellfish, or (b) a maximum acceptable 

dietary intake based on observations on survival, 

growth, or reproduction in a chronic wildlife 

feeding study or a long-term wildlife field study. 

If no maximum permissible tissue concentration is 

available, go to Section X because no Final 

Residue Value can be derived. 
B. ~ioconcentrationactors (BCFS) and 

bioaccumulation factors (BAFs) are quotients of 

the concentration of a material in one or more 

tissues of an aquatic organism divided by the 

average concentration in the solution in which the 

organism had been living. A BCF is intended to 

account Only for net uptake directly from water, 

and thus almosthastobemeasured ina laboratory 

test. Some uptake during the bioconcentration test 

might not bedirectly fromwater if the food sorbs 

some of the test material before it is eaten by 

the test organisms. A BAF is intended to account 

for net uptake from both food and water in a realworld situation. A BAF almost has to be measured 

in a field situation in which predators accumulate 

the material directly from water and by consuming 

prey that itself could have accumulated the 

material from both food and water. The BCF and 

BAF are probably similar for a material with a low 

BCF, but the BAF is probably higher than the BCF 

for materials with high BCFs. Although BCFs are 

not too difficult to determine, very few BAFs have 

been measured acceptably because it is necessary 

to make enough measurements of the concentration 

of the material in water to show that it was 

reasonably constant for a long enough period of 

time over the range of territory inhabited by the 
organisms. Because so few acceptable BAFs are 

available, only BCFs will be discussed further. 

However, if an acceptable BAF is available for a 

material, it should be used instead of any 

available BCFs. 

C. If a maximum permissible tissue concentration is 

available for a substance (e.g., parent material, 

parent material plus metabolites, etc.) , the 

tissue concentration used in the calculation of 

the BCF should be for the same substance. 

otherwise the tissue concentration used in the 

calculation of the BCF should be' that of the 

material and its metabolites which are 

structurally similar and are not much more soluble 

in water than the parent material. 

D. 1. A BCF should be used only if the test was 

flow-through, the BCF was calculated based on 

measured concentrations of the test material 

in tissue and in the test solution, and the 

exposure continued at least until either 

apparent steady-state or 28 days was reached. 

Steady-state is reached when the BCF does not 

change significantly over a period of time, 

such a 2 days or 16 percent of the length 

of the exposure, whichever is longer. The BCF 

used from a test should be the highest of (a) 

the apparent steady-state BCF, if apparent 

steady-state was reached, (b) the highest BCF 
obtained, if apparent steady-state was not 

reached, and (c) the projected steady-state 

BCF, if calculated. 

2. Whenever a BCF is determined for a lipophilic 

material, the percent lipids should also be 

determined in the tissue(s) for which the BCF 

was calculated. 

3. A BCF obtained from an exposure that 

adversely affected the test organisms may be 

used only if it is similar to a BCF obtained 

with unaffected organisms of the same species 

at lower concentrations that did not cause 

adverse effects. 

4. Because maximum permissible tissue 

concentrations are almost never based on dry 

weights, a BCF calculated using dry tissue 

weights must be converted to a wet tissue 

weight basis. If no conversion factor is 

reported with the BCF, multiply the dry weight 

BCF by 0.1 for plankton and by 0.2 for 

individual species of fishes and invertebrates 

~171. 

5. If more than one acceptable BCF is available 

for a species, the geometric mean of the 

available values should be used, except that 

if the BCFs are from different lengths of 

exposure and the BCF increases with length of 
exposure, the BCF for the longest exposure 

should be used. 

E. If enough pertinent data exist, several residue 

values can be calculated by dividing maximum 

permissible tissue concentrations by appropriate 

BCFs : 

1. For each available maximum acceptable dietary 

intake derived from a chronic feeding study or 

a long -term field study with wildlife, 

including birds and aquatic organisms, the 

appropriate BCF is based on the whole body of 

aquatic species which constitute' or represent 

a major portion of the diet of the tested 

wildlife species. 

2. For an FDA action level for fish or shellfish, 

the appropriate BCF is the highest geometric 

mean species BCF for the edible portion 

(muscle for decapods, muscle with or without 

skin for fishes, adductor muscle for scallops, 

and total soft tissue for other bivalve 

molluscs) of a consumed species. The highest 

species BCF is used because FDA action levels 

are applied on a species-by-species basis. 

F. For lipophilic materials, it might be possible to 

calculate additional residue values. Because the 

steady-state BCF for a lipophilic material Seems 

to be proportional to percent lipids from one 

tissue to another and from one species to another 
[18-20], extrapolations can be made from tested 

tissues or species to untested tissues or species 

an the basis of percent lipids. 

1. For each BCF for which the percent lipids is 

known for the same tissue for which the BCF 

was measured, normalize the BCF to a 1 

percent lipid basis by dividing the BCF by the 

percent lipids. This adjustment to a 1 

percent lipid basis is intended to make all 

the measured BCFs far a material comparable 

regardless of the species or tissue'with which 

the BCF was measured. 

2. calculate the geometric mean normalized BCF. 

Data for both saltwater and freshwater 

species should be used to determine the mean 

normalized BCF, unless the data show that the 

normalized BCFs are probably not similar, 

3. Calculate all possible residue values by 

dividing the available maximum permissible 

tissue concentrations by the mean normalized 

BCF and by the percent lipids values 

appropriate to the maximum permissible tissue 

concentrations, i.e., 

(maximum permissible tissue concentration) 

Residue value = (mean normalized BCF)(appropriate percent lipids) 

tissue concentration) Residue value = (mean 

normalized BCF) (appropriate percent lipids) 

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 11 

for freshwater criteria and 10 for 

saltwater criteria because FDA action 

levels are applied on a species-byspecies basis to commonly consumed 

species. The highest lipid contents in 

the edible portions of important consumed 

species are about 11 percent for both the 

freshwater chinook salmon and lake 

trout and about 10 percent for the 

saltwater Atlantic herring [21]. 

c. For a maximum acceptable dietary intake 

derived froma chronic feeding studyora 

long-term field study with wildlife, the 

appropriate percent lipids is that of an 

aquatic species or group of aquatic 

species which constitute a major portion 

of the diet of the wildlife species. 

G. The Final Residue Value is obtained by selecting 

the lowest of the available residue values. 
NOTE: In some cases the Final Residue Value will not be 

low enough. For example, a residue value 

calculated from an FDA action level will probably 

result in an average concentration in the edible 

portion of a fatty species that is at the action 

level. Some individual organisms, and possibly 

some species, will have residue concentrations 

higher than the mean value but no mechanism has 

been devised to provide appropriate additional 

protection. Also, some chronic feeding studies 

and long-term field studies with wildlife identify 

concentrations that cause adverse effects but do 

not identify concentrations which do not cause 

adverse effects; again, no mechanism has been 

devised to provide appropriate additional 

protection. These are some of the species and 

uses that are not protected at all times in all 

places. 

- x. other Data 

Pertinent information that could not be used in 

earlier sections might be available concerning adverse 

effects on aquatic organisms and their uses. The most 

important of these are data on cumulative and delayed 

toxicity, flavor impairment, reduction in survival, 

growth, or reproduction, or any other adverse effect 

that has been shown to be biologically important. 

Especially important are data for species for which no 

~~ 
other data are available. Data from behavioral, 

biochemical, physiological, microcosm, and field 

studies might also be available. Data might be 

available from tests conducted in unusual dilution 

water (see 1V.D and VI.D), from chronic tests in which 

the concentrations were not measured (see VI.B), from 

tests with previously exposed organisms (see 1I.F) , 

and from tests on formulated mixtures or emulsifiable 

concentrates (see 1I.D). Such data might affect a 

criterion if the data were obtained with an important 

species, the test concentrations were measured, and 

the endpoint was biologically important. ' 

_. 

XI. Criterion 

A. A criterion consists of two concentrations: the 

Criterion Maximum Concentration and the Criterion 

Continuous Concentration. 

B. The Criterion Maximum Concentration (CMC) is equal 

to one-half the Final Acute Value. 

C. The Criterion Continuous Concentration (CCC) is 

equal to the lowest of the Final Chronic Value, 

the Final Plant Value, and the Final Residue 

Value, unless other data (see Section X) show that 

a lower value should be used. If toxicity is 

related to a water quality characteristic, the 

Criterion continuous concentration is obtained 

from the Final Chronic Equation, the Final Plant 

Value, and the Final Residue Value by selecting 
the one, or the combination, that results in the 

lowest concentrations in the usual range of the 

water quality characteristic, unless other data 

(see Section X) show that a lower value should be 

used. 

D. Round [14] both the Criterion Maximum 

Concentration and the Criterion Continuous 

Concentration to two significant digits. 

E. The criterion is stated as: 

The procedures described in the Guidelines for 

Deriving Numerical National Water Quality Criteria 

for the Protection of Aquatic Organisms and Their 

Uses indicate that, except possibly where a 

locally important species is very sensitive, (1) 0 aquatic organisms and their uses should not be 

affected unacceptably if the 4-day average 

concentration of (2) does not exceed (3) ug/L more 

than once every 3 years on the average and if the 

1-hour average concentration does not exceed (4) 

ug/L more than once every 3 years on the average. 

where (1) = insert nlfreshwaternl or nrsaltwaternl 

(2) = insert name of material 

(3) = insert the Criterion Continuous 

Concentration 

(4) = insert the Criterion Maximum 

Concentration. 
Final Reviev 

A. The derivation of the criterion should be 

carefully reviewed by rechecking each step of the 

Guidelines. Items that should be especially 

checked are: 

1. If unpublished data are used, are they well 

documented? 

2. Are all required data available? 

3. Is the range of acute values for any species 

greater than a factor of lo? 

4. Is the range of Species Mean Acute Values for 

any genus greater than a factor of lo? 

5. Is there more than a factor of 10 difference 

between the 'four lowest Genus Mean Acute 

Values? 

6. Are any of the four lowest Genus Mean Acute 

Valuesquestionable? 

7. Is the Final Acute Value reasonable in 

comparison with the Species Mean Acute Values 

and Genus Mean Acute Values? 

8. For any commercially or recreationally 

important species, is the geometric mean of 

the acute values from flow-through tests in 

which the concentrations of test material were 

measured lower than the Final Acute Value? 

9. Are any of the chronic values questionable? 

0 
10. 

11. 

12 

13. 

14 * 

15. 

16. 

Are chronic values available for acutely 

sensitive species? 

Is the range of acute-chronic ratios greater 

than a factor of 10? 

Is the Final Chronic Value reasonable in 

comparison with the available acute and 

chronic data? 

Is the measured or predicted chronic value for 

any commercially or recreational ly important 

species below the Final Chronic Value? 

Are any of the other data important? 

Do any data look like they might be outliers? 

Are there any deviations from the Guidelines? 

Are they acceptable? 

B. 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 these Guidelines. 
APPENDIX B 
SUMHARY OF THE 1980 AQUATIC LIFE GUIDELINES 

The Guidelines for Deriving Water Quality Criteria for the 

Protection of Aquatic Life and its Uses were developed to 

describe an objective, internally consistent, and appropriate way 

of ensuring that water quality criteria for aquatic life would 

provide, on the average, a reasonable amount of protection. The 

resulting criteria are not intended to provide 100 percent 

protection of all species and all uses of aquatic life all of the 

time, but they are intended to protect most species in a 

balanced, healthy aquatic community. 

Minimum data requirements are identified in four areas; 

acute toxicity to animals (eight data points), chronic toxicity 

to animals (three data points), toxicity to plants, and residues. 

Data on acute toxicity are needed for a variety of fish and 

invertebrate species and are used to derive a Final Acute Value. 

By taking into account the number and relative sensitivities of 

the tested species, the Final Acute Value is designed to protect 

most, but not necessarily all, of the tested and untested 

species. 

Data on chronic toxicity to animals can be used to derive a 

Final Chronic Value by two different means. If chronic values 

are available for a specified number and array of species, a 

Final Chronic Value can be calculated directly. If not, an 

acute-chronic ratio is derived and then used with the Final Acute 

Value to obtain the Final Chronic Value. 

The Final Plant Value is obtained by selecting the lowest 

plant toxicity value based on measured concentrations. 
The Final Residue Value is intended to protect wildlife which 

consume aquatic organisms and the marketability of aquatic 

organisms. Protection of the marketability of aquatic organisms 

is, in actuality, protection of a use of that water body 

(commercial fishery). Two kinds of data are necessary to 

calculate the Final Residue Value: a bioconcentration factor 

(BCF) and a maximum permissible tissue concentration, which can 

be an FDAaction level or canbetheresultof achronicwildlife 

feeding study. For lipid -soluble pollutants, the BCF is 

normalized for percent lipids and then the Final Residue Value is 

calculated by dividing the maximum permissible tissue 

concentration by the normalized BCF and by an appropriate percent 

lipid value. BCFs are normalized for percent lipids since the 

BCF measured for any individual aquatic species is generally 

proportional to the percent lipids in that species. a 

If sufficient data are available to demonstrate that one or 

more of the final values should be related to a water quality 

characteristic, such as salinity, hardness, or suspended sol ids, 

the final value(s) are expressed as a function of that 

characteristic. 

After the four final values (Final Acute Value, Final Chronic 

Value, Final Plant Value, and Final Residue Value) have been 

obtained, the criterion is established with the Final Acute value 

becoming the maximum value and the lowest of the other three 

values becoming the 24-hour average value. All of the data used 

to calculate the four final values and any additional pertinent 

information are then reviewed to determine if the criterion is 

reasonable. If sound scientific evidence indicates that the 
criterion should be raised or lowered, appropriate changes are 

made as necessary. 

The November 28, 1980, Guidelines have been revised from the 

earlier published versions (43 FR 21506, May 18, 1978; 43 FR 

29028, July 5, 1978; 44 FR 15926, March 15, 1979). Details have 

been added in many places and the concept of a minimum data base 

has been incorporated. In addition, three adjustment factors and 

the species sensitivity factor have been deleted. These 

modifications were the result of the Agency's analysis of public 

comments and comments received from the Science Advisory Board on 

earlier versions of the Guidelines. These comments and the 

Resultant modifications are addressed fully in Appendix D to this 

notice. 

Criteria for the Protection of Human Health 

Interpretation of the Human Health Criteria 

0 

The human health criteria issued today are summarized in 

Appendix A of this Federal Register notice. Criteria for the 

protection of human health are based on their carcinogenic, 

toxic, or organoleptic (taste and odor) properties. The meanings 

and practical uses of the criteria values are distinctly 

different depending on the properties on which they are based. 

The objective of the health assessment portions of the 

criteria documents is to estimate ambient water concentrations 

which, in the case of noncarcinogens, prevent adverse health 

effects in humans, and in the case of suspect or proven 

carcinogens, represent various levels of incremental cancer risk. 
Health assessments typically contain discussions of four 

elements: exposure, pharmacokinetics, toxic effects, and 

criterion formulation. 

The exposure section summarizes information on exposure 

routes: ingestion directly from water, indirectly from 

consumption of aquatic organisms found in ambient water, other 

dietary sources, inhalation, and dermal contact. Exposure 

assumptions are used to derive human health criteria. Most 

criteria are based solely on exposure from consumption of water 

containing a specified concentration of a toxic pollutant and 

through consumption of aquatic organisms which are assumed to 

have bioconcentrated pollutants from the water in which they 

live. Other multimedia routes of exposure such as air, 

nonaquatic diet, or dermal are not factored into the criterion 

formulation for the vast majority of pollutants because of lack 

of data. The criteria are calculated using the combined aquatic 

exposure pathway and also using the aquatic organism ingestion 

exposure route alone. In criteria reflecting both the water 

consumption and aquatic organism ingestion routes of exposure, 

the relative exposure contribution varies with the propensity of 

a pollutant to bioconcentrate, with the consumption of aquatic 

organisms becoming more important as the bioconcentration factor 

(BCF) increases. As additional information on total exposure is 

assembled for pollutants for which criteria reflect only the two 

specified aquatic exposure routes, adjustments in water 

concentration values may be made. The demonstration of 

significantly different exposure patterns will become an element 
of a process to adapt/modify human health-based criteria to local 

conditions, somewhat analogous to the aquatic life criteria 

modification process discussed previously. It is anticipated 

that States at their discretion will be able to set appropriate 

human health criteria based on this process. 

0 

Specific health-based criteria are developed only if a weight 

of evidence supports the occurrence of the toxic effect and if 

dose/response data exist from which criteria can be estimated. 

The pharmacokinteics section reviews data on absorption, 

distribution, metabolism, and excretion to assess the biochemical 

fate of the compounds in the human and animal system. The toxic 

effects section reviews data on acute, subacute, and chronic 

toxicity, synergistic and antagonistic effects, and specific 

information on mutagenicity, teratogenicity, and carcinogenicity. 

From this review, the toxic effect to be protected against is 

identified taking into account the quality, quantity, and weight 

of evidence characteristic of the data. The criterion 

formulation section reviews the highlights of the text and 

specifies a rationale for criterion development and the 

mathematical derivation of the criterion number. 

Within the limitations of time and resources, current 

published information of significance was incorporated into the 

human health assessments. Review articles nad reports were used 

for data evaluation and synthesis. Scientific judgment was 

exercised in reviewing and evaluating the data in each criteria 

document and in identifying the adverse effects for which 0 >. protective criteria were published. 
Criteria for suspect or proven carcinogens are presented as 

concentrations in water associated with a range of incremental 

cancer risks to man. Criteria for noncarcinogens represent 

levels at which exposure to a single chemical is not anticipated 

to produce adverse effects in man. In a few cases, organoleptic 

(taste and odor) data form the basis for the criterion. While 

this type of criterion does not represent a value which directly 

affects human health, it is presented as an estimate of the level 

of a pollutant that will not produce unpleasant taste or odor 

either directly from water consumption or indirectly by 

consumption of aquatic organisms found in ambient waters. A 

criterion developed in this manner is judged to be as useful as 

other types of criteria in protecting designated water uses. In 

addition, where data are available, toxicity-based criteria are 

also presented for pollutants with derived organoleptic criteria. 

The choice of criteria used in water quality standards for these 

pcllutants will depend upon the designated use to be protected. 

In the case of a multiple use water body, the criterion 

protecting the most sensitive use will be applied. Finally, for 

several pollutants no criteria are recommended because 

insufficient information is available for quantitative criterion 

formulation. 

Risk Extrapolation 

Because methods do not exist to establish the presence of a 

threshold for carcinogenic effects, EPAIs policy is that there is 

no scientific basis for estimating "safe" levels for carcinogens. 

The criteria for carcinogens, therefore, state that the 
recommended concentration for maximum protection of human health 

is zero. In addition, the Agency has presented a range of 

concentrations corresponding to incremental cancer risks of 

to (one additional case of cancer in populations ranging 

from 10 million to 100,000, respectively). Other concentrations 

representing different risk levels may be calculated by use of 

the Guidelines. The risk estimate range is presented for 

information purposes and does not represent an Agency judgment on 

a "acceptable" risk level. 

Summary of the Human Health Guidelines 

The health assessments and corresponding criteria were 

derived based on Guidelines and Methodology Used in the 

Preparation of Health Effect Assessment Chapters of the Consent 

Decree Water Criteria Documents (the Guidelines ) developed by 0 

EPA'S Office of Research and Development. The estimation of 

health risk associated with human exposure to environmental 

pollutants requires predicting the effect of low doses for up to 

a lifetime in duration. A combination of epidemiological and 

animal dose/response data is considered the preferred basis for 

quantitative criterion derivation. 

No-effect (noncarcinogen) or specified risk (carcinogen) 

concentrations were estimated by extrapolation from animal 

toxicity or human epidemiology studies using the following basic 

exposure assumptions: a 70-kilogram male person (Report of the 

Task Group on Reference Man, International Commission for 

Radiation Protection, November 23, 1957) as the exposed 0 individual; the average daily consumption of freshwater and 
estuarine fish and shellfish products equal to 6.5 grams/day; and 

the average ingestion of 2 liters/day of water (Drinking Water 

and Health, National Academy of Sciences, National Research 

Council, 1977). Criteria based on these assumptions are 

estimated to be protective of an adult male who experiences 

average exposure conditions. 

Two basic methods were used to formulate health criteria, 

depending on whether the prominent adverse effect was cancer or 

other toxic manifestations. The €01 lowing sections detail these 

methods. 

Carcinogens 

Extrapolation of cancer responses from high to low doses and 

subsequent risk estimation from animal data are performed using a 

linearized multi-stage model. This procedure is flexible enough 

to fit all monotonically-increasing dose response data, since it 

incorporates several adjustable parameters. The multi-stage 

model is a linear nonthreshold model as was the "one-hit" model 

original 1 y used in the proposed criteria documents. The 1 inear 

nonthreshold concept has been endorsed by the four agencies in 

the Interagency Regulatory Liaison Group and is less likely to 

underestimate risk at the low doses typical of environmental 

exposure than other models that could be used. Because of the 

uncertainties associated with dose response, animal-to-human 

extrapolation, and other unknown factors; because of the use of 

average consumptions; and because of the serious public health 

consequences that could result if risks were underestimated, EPA 

believes that it is prudent to use conservative methods to 
estimate risk in the water quality criteria program. The 

linearized multistage model is more systematic and invokes fewer 

arbitrary assumptions than the “one-hit” procedure previously 

used. 

0 

It should be noted that extrapolation models provide 

estimates of risk since a variety of assumptions are built into 

any model. Models using widely different assumptions may produce 

estimates ranging over several orders of magnitude. Since there 

is at present no way to demonstrate the scientific validity of 

any model, the use of risk extrapolation models is a subject of 

debate in the scientific community. However, risk extrapolation 

is generally recognized as the only tool available at this time 

€or estimating the magnitude of health hazards associated with 

nonthreshold toxicants and has been endorsed by numerous Federal 

agencies and scientific organizations, including EPA‘s Carcinogen 

Assessment Group, the National Academy of Sciences, and the 

Interagency Regulatory Liaison Group, as a useful means of 

assessing the risks of exposure to various carcinogenic 

pollutants. 

Noncarcinogens 

Health criteria based on toxic effects of pollutants other 

than carcinogenicity are estimates of concentrations which are 

not expected to produce adverse effects in humans. They are 

based upon Acceptable Daily Intake (ADI) levels and are generally 

derived using no-observed-adverse-ef fect-level data from animal 

studies although human data are used wherever available. The AD1 0 

is calculated using safety factors to account for uncertainties 
inherent in extrapolation from animal to man. In accordance With 

the National Research Council recommendations (Drinking Water and 

Health, National Academy of Sciences, National Research Council, 

1977), safety factors of 10, 100, or 1,000 are used, depending on 

the quality and quantity of data. In some instances 

extrapolations are made from inhalation studies or limits to 

approximate a human response from ingestion using the StokingerWoodward model (Journal of American Water Works Association, 

1958). Calculations of criteria from ADIS are made using the 

standard exposure assumptions (2 liters of water, 6.5 grams of 

edible aquatic products, and an average body weight of 70 kg). 
APPENDIX C 

. 
- THE PHILOSOPHY --- OF TRE 1976 WATER QUALITY CRITERIA 

Water quality criteria specify concentrations of water 

constituents which, if not exceeded, are expected to support an 

organic ecosystem suitable for the higher uses of water. Such 

criteria are derived from scientific facts obtained from 

experimental or L~J g&gg observations that depict organic 

responses to a defined stimulus or material under identifiable or 

regulated environmental conditions for a specified time period. 

Water quality criteria are not intended to offer the same 

degree of strategy for survival and propagation at all times to 

all organisms within a given ecosystem. They are intended not 

only to protect essential and significant life in water and the 

direct users ofwater,butalsotoprotect life that isdependent 

on life in water for its existence, or that may consume 

intentionally or unintentionally any edible portion of such life. 

The criteria levels for domestic water supply incorporate 

available data for human health protection. Such values are 

different from the criteria levels necessary for protection of 

aquatic life. The Agency's interim primary drinking water 

regulations (40 Federal Register 59566 December 24, 1975), as 

required by the Safe Drinking Water Act (42 U.S.C. 300f, et 

seq.) , incorporate applicable domestic water supply criteria. 

Where pollutants are identified in both the quality criteria for 

domestic water supply and the Drinking Water Standards, the 

concentration levels are identical. Water treatment may not 

significantly affect the removal of certain pollutants. 
What is essential and significant life in water? Do Daphnia 

or stonefly nymphs qualify as such life? Why does 1/100th of a 

concentration that is lethal to 50 percent of the test organisms 

(LC50) constitute a criterion in some instances, whereas 1/2 or 

l/lOth of some effect levels constitutes a criterion in other 

instances? These are questions that often are asked of those 

who undertake the task of criteria formulation. 

The universe of organisms composing life in water is great in 

both kinds and numbers. As in the human population, 

physiological variability exists among individuals of the same 

species in response to a given stimulus. A much greater response 

variation exists among species of aquatic organisms. Thus, 

aquatic organisms do not exhibit the same degree of harm, 

individually or by species, from a given concentration of a 

toxicant or potential toxicant within the environment. In 

establishing a level or concentration of a quality constituent as 

a criterion it is necessary to ensure a reasonable degree of 

safety for those more sensitive species that are important to the 

functioning of the aquatic ecosystem even though data on the 

response of such species to the quality constituent under 

consideration may not be available. The aquatic food web is an 

intricate relationship of predator and prey organisms. A water 

constituent that may in some way destroy or eliminate an 

important segment of that food web would, in all likelihood, 

destroy or seriously impair other organisms associated with it. 

1 

0 Although experimentation relating to the effects of 

particular substances under control led conditions began in the 
early 19OO's, the effects of any substance on more than a few of 

the vast number of aquatic organisms have not been investigated. 

Certain test animals have been selected by investigators for 

intensive investigation because of their importance to man, their 

availability to the researcher, and their physiological responses 

to the laboratory environment. As general indicators of organism 

responses such test organisms are representative of the expected 

results for other associated organisms. In this context Daphnia 

or stoneflies or other associated organisms indicate the general 

levels of toxicity to be expected among untested species. In 

addition, test organisms are themselves vital 1 inks within the 

food web that results in the fish population in a particular 

waterway. 

The ideal data base for criteria development would consist of 

information on a large percentage of aquatic species and would 

show the community response to a range of concentrations for a 

tested constituent during a long time period. This information 

is not available but investigators are beginning to derive such 

information for a few water constituents. Where only 96-hour 

bioassay data are available, judgmental prudence dictates that a 

substantial safety factor be employed to protect all life stages 

of the test organism in waters of varying quality, as well as 

associated organisms within the aquatic environment that have not 

been tested and that may be more sensitive to the test 

constituent. Application factors have been used to provide the 

degree of protection required. Safe levels for certain 

chlorinated hydrocarbons and certain heavy metals were estimated 

by applying an 0.01 application factor to the 96-hour LC50 value 

a 
for sensitive aquatic organisms. Flow-through bioassays have 

been conducted for some test indicator organisms over a 

substantial period of their life history. In a few other cases, 

information is available for the organism's natural life or for 

more than one generation of the species. Such data may indicate 

a minimal effect level, as well as a no-effect level. 

The word '*criterion*I should not be used interchangeably with 

or as a synonym for the word '*standard.** The word '*criterion" 

represents a constituent concentration or level associated with a 

degree of environmental effect upon which scientific judgment may 

be based. As it is currently associated with the water 

environment it has come to mean a designated concentration of a 

constituent that, when not exceeded, will protect an organism, 

an organism community, or a prescribed water use or quality with 

an adequate degree of safety. A criterion, in some cases, may be 

a narrative statement instead of a constituent concentration. on 

the other hand, a standard connotes a legal entity for a 

particular reach of waterway or €or an effluent. A water quality 

standard may use a water quality criterion as a basis for 

regulation or enforcement, but the standard may differ from a 

criterion because of prevailing local natural conditions, such as 

naturally occurring organic acids, or because of the importance 

of a particular waterway, economic considerations, or the degree 

of safety to a particular ecosystem that may be desired. 

Toxicity to aquatic life generally is expressed in terms of 

acute (short term) or chronic (long-term) effects. Acute 

toxicity refers to effects occurring in a short time period: 
often death is the end point. Acute toxicity can be expressed as 

the lethal concentration for a stated percentage of organisms 

tested, or the reciprocal, which is the tolerance limit of a 

percentage of surviving organisms, Acute toxicity for aquatic 

organisms generally has been expressed for 24 to 96-hOur 

exposures. 

chronic toxicity refers to effects through an extended time 

period. Chronic toxicity may be expressed in terms of an 

observation period equal to the lifetime of an organism or to the 

time span of more than one generation. Some chronic effects may 

be reversible, but most are not. 

Chronic effects often occur in the species population rather 

than in the individual. or the sperm 

does not remain viable, the species would be eliminated from an 

ecosystem because of reproductive failure. Physiological stress 

may make a species less competitive with others and may result in 

a gradual population decline or absence from an area. The 

elimination of a microcrustacean that serves as a vital food 

during the larval period of a fish's life could result ultimately 

in the elimination of the fish from an area. The phenomenon of 

bioaccumulation of certain materials may result in chronic 

toxicity to the ultimate consumer in a rood chain. Thus, fish 

may mobilize lethal toxicants from their fatty tissues during 

periods of physiological stress. Egg shells of predatory birds 

may be weakened to a point of destruction in the nest. Bird 

chick embryos may have increased mortality rates. There may be a 

hazard to the health of man if aquatic organisms with toxic 

residues are consumed. 

If eggs fail to develop 
The fact that living systems, i.e., individuals, populations, 

species, and ecosystems, can take up, accumulate, and 

bioconcentrate manmade and natural toxicants is well documented. 

In aquatic systems biota are exposed directly to pollutant 

toxicants through submersion in a relatively efficient solvent 

(water) and are exposed indirectly through food webs and other 

biological, chemical, and physical interactions. Initial 

toxicant levels, if not immediately toxic and damaging, may 

accumulate in the biota or sediment over time and increase to 

levels that are lethal or sublethally damaging to aquatic 

organisms or to consumers of these organisms. Water quality 

criteria reflect a knowledge of the capacity for environmental 

accumulation, persistence, and effects of specific toxicants in 

specific aquatic systems. 

Ions of toxic materials frequently cause adverse effects 

because they pass through the semipermeable membranes of an 

organism. Molecular diffusion through membranes may occur for 

some compounds such as pesticides, polychlorinated biphenyls, 

and other toxicants. Some materials may not pass through 

membranes in their natural or waste-discharged state, but in 

water they may be converted to states that have increased ability 

to affect organisms. For example, certain microorganisms can 

methylate mercury, thus producing a material that more readily 

enters physiological systems. Some materials may have multiple 

effects: for example, an iron salt may not be toxic; an iron floc 

or gel may be an irritant or clog fish gills to effect 

asphyxiation; iron at low concentrations can be a trace nutrient 
but at high concentrations it can be a toxicant. Materials also 

can affect organisms if their metabolic byproducts cannot be 

excreted. Unless otherwise stated, criteria are based on the 

total concentration of the substance because an ecosystem can 

produce chemical, physical, and biological changes that may be 

detrimental to organisms living in or using the water. 

In prescribing water quality criteria, certain fundamental 

principles dominate the reasoning process. In establishing a 

level or concentration as a criterion for a given constituent it 

was assumed that other factors within the aquatic environment 

are acceptable to maintain the integrity of the water. 

Interrelationships and interactions among organisms and their 

environment, as well as the interrelationships of sediments and 

the constituents they contain to the water above, are recognized 0 as fact. 

Antagonistic and synergistic reactions among many quality 

constituents in water also are recognized as fact. The precise 

definition of such reactions and their relative effects on 

particular segments of aquatic life have not been identified with 

scientific precision. Historically much of the data to support 

criteria development was of an ambient concentration-organism 

response nature. Recently, data are becoming available on longterm chronic effects on particular species. Studies now 

determine carcinogenic, teratogenic, and other insidious effects 

of toxic materials. 

Some unpolluted waters in the Nation may exceed designated 

criteria for particular constituents. There is variability in 

the natural quality of water and certain organisms become adapted 
to that quality, which may be considered extreme in other areas. 

Likewise, it is recognized that a single criterion cannot 

identify minimal quality for the protection of the integrity of 

water for every aquatic ecosystem in the Nation. To provide an 

adequate degree of safety to protect against long-term effects 

may result in a criterion that cannot be detected with present 

analytical tools. In some cases, a mass balance calculation can 

provide a means of assurance that the integrity of the waterway 

is not being degraded. 

Water quality criteria do not have direct regulatory impact, 

but they form the basis for judgment in several Environmental 

Protection Agency programs that are derived from water quality 

considerations. For example, water qual ity standards developed 

by the States under section 303 of the Act and approved by EPA 

are to be based on the water quality criteria, appropriately 

modified to take account of local conditions. The local 

conditions to be considered include actual and projected uses of 

the water, natural background levels of particular constituents, 

the presence or absence of sensitive important species, 

characteristics of the local biological community, temperature 

and weather, flow characteristics, and synergistic or 

antagonistic effects of combinations of pollutants. 

Similarly, by providing a judgment on desirable levels of 

ambient water qual ity, water quality criteria are the starting 

point in deriving toxic pollutant effluent standards pursuant to 

section 307(a) of the Act. Other EPA programs that use water 

qual ity criteria involve drinking water standards, the ocean 
dumping program, designation of hazardous substances, dredge 

spoil criteria development, removal of in-place toxic materials, 

thermal pollution, and pesticide registration. 

To provide the water resource protection for which they are 

designed, quality criteria should apply to virtually all of the 

Nation's navigable waters with modifications for local conditions 

as needed. To violate quality criteria for any substantial 

length of time or in any substantial portion of a waterway may 

result in an adverse affect on aquatic life and perhaps a hazard 

to man or other consumers of aquatic life. 

Quality criteria have been designed -to provide long-term 

protection. Thus, they may provide a basis for effluent 

standards, but it is not intended that criteria values become 

effluent standards. It is recognized that certain substances may a be applied to the aquatic environment with the concurrence of a 

governmental agency for the precise purpose of controlling or 

managing a portion of the aquatic ecosystem: aquatic herbicides 

and piscicides are examples of such substances. For such 

occurrences, criteria obviously do not apply. It is recognized 

further that pesticides applied according to official label 

instructions to agricultural and forest lands may be washed to a 

receiving waterway by a torrential rainstorm. Under such 

conditions it is believed that such diffuse source inflows should 

receive consideration similar to that of a discrete effluent 

discharge and that in such instances the criteria should be 

applied to the principal portion of the waterway rather than to e -. , that peripheral portion receiving the diffuse inflow. 
The format for presenting water quality criteria includes a 

concise statement of the dominant criterion or criteria for a 

particular constituent followed by a narrative introduction, a 

rationale that includes justification for the designated 

criterion or criteria, and a listing of the references cited 

within the rationale. An effort has been made to restrict 

supporting data to those which either have been published or are 

in press awaiting publication. A particular constituent may have 

more than one criterion to ensure more than one water use or 

condition, i.e., hard or soft water where applicable, suitability 

as a drinking water supply source, protection of human health 

when edible portions of selected biota are consumed, provision 

for recreational bathing or waterskiing, and permitting an 

appropriate factor of safety to ensure protection for essential 

warm-or coldwater associated biota. 

Criteria are presented €or those substances that may occur in 

water where data indicate the potential for harm to aquatic life, 

or to water users, or to the consumers of the water or aquatic 

life. Presented criteria do not represent an all-inclusive list 

of constituent contaminants. omissions from criteria should not 

be construed to mean that an omitted quality constituent is 

either unimportant or non-hazardous. 
BIBLIOGRAPHY 

Adelman, I.R. and L.L. Smith, 1970. Effect of hydrogen sulfide 

on northern pike eggs and sac fry. 

99:501. 

Trans. Amer. Fish. SOC., 

Agriculture Handbook No. 60, 1954. Diagnosis and improvement of 

saline and alkali soils. L.A. Richards, editor, U.S. Govern- ment Printing Office, Washington, D.C. 

Anderson, B.G., 1960. The toxicity of organic insecticides to 

Daphnia. Second Seminar on Biol. Problems in Water Pollution. 

Robert A. Taft Sanitary Engineering Center Technical Report 

W60-3, Cincinnati, Ohio. 

Bahner, L.H. and D.R. Nimmo, 1974. Methods to assess effects of 

combinations of toxicants, salinity, and temperature on estumarine animals. Proc. 9th Am. Conf. on Trace Substances in 

Env. Health, Univ. Miss.?olumbia, Mo. 

Ballentine, R.X. and F.W. Kittrell, 1968. Observations of fecal 

colifonns in several recent stream pollution studies. 

ings of the Symposium on Fecal Coliform Bacteria in Water and 

Wastewater, May 21-22, 1968, Bureau of Sanitary Engineering, 

California State Department of Public Health. 

Bell, H.L., 1971. Effect of low pH on the survival and emergence 

of aquatic insects. Water Res., 5:313. 

Bellan, et al., 1972. The sublethal effects of a detergent on 

the reproduction, development, and settlement in the polychaetous annelid Capitella capitata. Marine Biology, 14:183. 

ProceedBender, M.E., 1969. The toxicity of the hydrolysis and breakdown 

products of malathion to the fathead minnow (PimephalE 

promelas, Rafinesque). Water Res., 3:571. 

methyl parathion, parathion and azinphosmethyl in mice and 

fish: Onset and recovery of inhibition. Bull. Environ. 

Contam. Toxicol., 12:117. 

Berg, G., 1974. Reassessment of the virus problem in sewage and 

in surface and renovated waters. Sixth International Water 

poll. Res. California. Pergamon Press. 

Bigqar, J.W. and M. Fireman, 1960. Boron abosrption and release 

by soils. 

Billard, R. and dekinkelin, 1970. Sterilization of the testicles 

of guppies by means of non-lethal doses of parathion. 

D Hydrobiologie, 9(1):91. 

Benke, G.M. and S.D. Murphy, 1974. Anticholinesterase action of 

Soil Sci. SOC. Amer. Proc., 24:115 

Annales 
Black, E.C., 1953. Upper lethal temperatures of some British 

Columbia freshwater fishes. Jour. Fish. Res. Bd. Can., 10:196 

Blumer, M., 1970. Oil contamination and the living resources of 

the sea. F.A.O. Tech. Conf. Rome. FIR:MP/7O/R-lIllP. 

Bollard, E.G. and G.W. Butler, 1966. Mineral nutrition of 

plants. Ann. Rev. Plant Physiol., 17:77. 

Bonde, G.J., 1966. Bacteriological methods for the estimation of 

water pollution. Hlth. Lab. Sci. 3:124. 

Bonde, G.J., 1974. Bacterial indicators of sewage pollution. 

International Symposium on Discharge of Sewage from Sea Out- falls. Pergamon Press. 

Bookhout, C.G. - et -.I a1 1973. Effects of mirex on the larval 

development of two crabs. Water, Air, and Soil Pollution, 

1:165. 

channel catfish. Trans. Amer. Fish. SOC., 96:31. 

mirex in selected estuaries of South Carolina, 1969-71. Pesti- cides Monit. Jour., 7:6. 

an reproduction of fish exposed to gas supersaturated water. 

Unpublished report. U.S. Environmental Protection Agency, 

Western Fish Toxicology Station, Corvallis, Oregon. 

Boyd, C.E. and D.E. Ferguson, 1964. Susceptibility and 

resistance of mosquitofish to several insecticides. Econ. 

Entomo 1. , 57: 430. 

Boyle, H.W., 1967. Taste/odor contamination of fish from the 

Ohio River. Fed. Water Poll. Cont. Admin., Cincinnati, Ohio. 

Bradford, G.R., 1966. Boron [toxicity, indicator plants] , in 

diagnostic criteria for plants and soils. H.D. Chapman, Ed., 

University of California, Division of Agricultural Science, 

Berkeley, p. 33. 

development and weight of sockeye salmon embryos and alevins. 

International Pacific Salmon Fisheries Commission, Progress 

Report No. 12, pp. 1-26, mimeo. 

of speckled trout. Trans. Amer. Fish. SOC., 70:397. 

of fishes. Quarterly Rev. Biol., 31:75. 

Boon, C.W. and B.J. Follis, 1967. Effects of hydrogen sulfide on 

Borthwick, P.W., et al., 1973. Accumulation and movement of 

Bouck, G.R., et al., 1975. Mortality, saltwater adaption 

Brannon, E.L., 1965. The influence of physical factors on the 

Brett, J.R., 1941. Tempering versurs acclimation in the planting 

Brett, J.R., 1956. Some principals in the thermal requirements 
Brett, J.R., 1960. Thermal requirementsof f ish--three decades of 

study, 1940-1970. In: Biological problems in water pollution. 

C.M. Tarzwell (ed.) Dept. of Health, Education and Welfare, 

Public Health Service. 

Brinley, F.J. , 1944. Biological Studies. House Document 266, 

78th Congress, 1st session: Part 11, Supplement F. pp. 1275- 

1353. 

0 

Brungs, W.A., 1971. Chronic effects of low dissolved oxygen 

concentrations on fathead minnow (Pimephales promelas). - J. Fish. Res. Bd. Canada. 28:1119-1123. 

taste in domestic water. Jour. Amer. Water Works Assn., 

61:575. 

Bugbee, S.L. and C.M. Walter, 1973. The response of macroinvertebrates to gasoline pollution in a mountain stream. In: 

Prevention and control of oil spills, proceedings of symposium 

March 13-17, Washington, D.C., p. 725. 

insecticides to resistant and susceptible mosquitofish in 

static and flowing solutions. , Mosquito News, 29(1):96. 

Burson, B., 1938. Seasonal temperature variations in relation 

to water treatment. Jour. Amer. Water Works Assn., 30:793. 

Butler, P.A., 1963. Commercial fisheries investigations. In: 

pesticide wildlife Studies during 1961 and 1962. U.S. Fish. 

Wildl. Serv. Circ. 167, Washington, D.C. 

organisms. In: A Guide to Marine Pollution, E.D. Goldberg, 

ed. Gordon and Beach, NY. 

Calabrese, A., 1969. Individual and combined effects of salinity 

and temperature on embryos and larvae of the coot clam, Mulinia 

- lateralis (say), Biol. Bull. 137, 3:417. 

embryos of the american oyster, Crassotrea virginica. Marine 

Biol. , 18: 162. 

Camp, T.R., 1963. Water and its impurities. Reinhold Publishing 

Corp, New York, New York. 

Capurro, L.R.A., 1970. Oceanography for practicing engineers. 

Barnes and Noble Inc., New York. 

Carlson, A.R., et al., 1974. Effects of lowered dissolved oxygen 

concentrations on channel catfish (Ictalurus embryos and larve. 

Trans. Amer. Fish. SOC. 103:623-626. 

Bruvold, W.H., et ,al., 1969. Consumer assessment of mineral 

Burke, W.D. and D.E. Ferguson, 1969. Toxicities of four 

0 

Butler, P., et al., 1972. Test, monitoring and indicator 

Calabrese, A., et al., 1973. The toxicity of heavy metals to 
Chin, E., 1961. A trawl study of an estuarine nursery area in 

Galveston Bay with particular reference to penaeid shrimp. 

P~.D. Dissertation, University of Washington. 

significance of lead isotopes in pelagic sediments. 

Cosmochim. Acta., 26:263. 

Clay, A., et a1 ., 1975. Experimental induction of gas bubble 

disease in menhaden. Presented a the American Fisheries 

Society, September, 1975, Las Vegas, Nevada. New England 

Aquarium, Boston, Mass. 

Coble, D.W., 1961. Influence of water exchange and dissolved 

oxygen in redds on survival of steelhead trout embryos, Trans. 

Amer. Fish. SOC. 90~469-474. 

Cooper, A.C., 1965. The effect of transported stream sediments 

on the survival of sockeye and pink salmon eggs and alevin. 

International Pacific Salmon Fisheries Commission, Bulletin 

Chow, T.J., and C.C. Patterson, 1962. The occurrences and 

Geochim. 

18 : 1-71. 

Coppage, D.L., 1972. Organophosphate pesticides: specific level 

of brain ACHE inhibition related to death in sheepshead 

minnows. Trans. Amer. Fish. SOC., 101:534. 

Coppage, D.L. and T.W. Duke, 1971. Effects of pesticides in 

estuaries along the Gulf and Southeast Atlantic Coasts. In: 

Proceedings of the 2nd Gulf Coast Conference on Mosquito 

Suppression and Wildlife Management (C.H. Schmidt, ed.) 

National Mosquito Control-Fish and Wildlife Management 

Coordinating Committee, Washington, D.C. 

Coppage, D.L. and E. Matthews, 1974. Short-term effects of 

organophosphate pesticides on cholinesterases of estuarine 

fishes and pink shrimp. Bull. Environ. Contam. Toxicol., 

11:483. 

inhibition in fish as a diagnosis of environmental poisoning by 

malathion, 0, 0-dimethyl S-(1,l-dicarbethoxy-ethyl) phosphorodithioate. Pesticide Biochemistry and Physiology (in press). 

temperatures on larval development in the mud crab, 

Rhitropanopeus harrisii. In: Fourth European Marine Biology 

Symposium. D.I. Crisp (ed.) , Cambridge University Press, 

London. 

review of the lterature of 1967. Jour. Water Poll. Cont. Fed., 

40:1047. 

Coutant, C.C., 1969. Thermal pollution effects: A review of the 

Coppage, D.L. .et al., 1975. Brain acteylcholinesterase 

Costlow, Jr., J.D. and C.G. Bookhout, 1971. The effect of cyclic 

Coutant, C.C., 1968. Thermal pollution--biologicaly effects: A 

literature of 1968. Jour. Water Poll. Cont. Fed., 41:1036. 
Coutant, C.C., 1970. Thermal pollution effects: A review of the 

literature of 1970. Thermal pollution effects: A review of 

the literature of 1969. Jour. Water Poll. Cont. Fed., 42:1025. 

Coutant, C.C., 1971. Thermal pollution effects: A review of the 

literature of 1970. Jour. Water Poll. Cont. Fed., 43:1292. 

Coutant, C.C., 1972. Biological aspects of thermal pollution, 

11. Scientific basis for water temperature standards at power 

plants. CRC Critical Rev. in Environ. Cont., 3:1. 

Coutant, C.C., 1975. Temperature selection by fish--a factor in 

power plant impact assessments. In: Symposium on the physical 

and biological effects on the environment of cooling systems 

and thermal discharges at nuclear power stations. Int'l Atomic 

Energy Agency (In press). 

Coutant, C.C., and C.P. Goodyear, 1972. Water pollution-thermal 

pollution: A review of the literature of 1971. Jour. Water 

Poll. Cont. Fed., 45:1250. 

Coutant, C.C., and H.A. Pfuderer, 1973. Water pollution-thermal 

effects: A review of the literature of 1973. Jour. Water 

Poll. Cont. Fed., 45:1331. 

effects: A review of the literature of 1974. Jour. Water 

Poll. Cont. Fed., 45:1476. 

marine animals. Nature, 179:1138. 

Coutant, C.C., and H.A. Pfuderer, 1974. Water pollution-thermal 

Crisp, D.J., 1957. Effect of low temperature on the breeding of 

Datta, N. and J. Olearte, 1974. R-Factors in strains of 

Salmonella typhi and Shigella dysenteri I isolated 

during efi7emic.s in Mexico: 

- 

Classification by comrJabilitv. 

Antimicrobial Agents and Chemotherapy 5:310. 

embryonic development of clams and oysters and on survival and 

grouwth of the larvae. U.S. Fish and Wildlife Service, Fishery 

Bulletin, 67:393. 

- - - 

Davis, H.C. and H. Hindu, 1969. Effects of pesticides on 

Dawlev, E.M. and W.J. Ebel. 1975. Effectw of various ~~ 

concentrations of dissolved atmospheric gas on juvenile chinook 

salmon, Oncorhynchus tsawytscha, and steelhead trout. Fish. 

Bull., U.S. (In press) 

Dawley, E., et al., 1975. Bioassays of total dissolved gas 

pressure. (Unpublished report. National Marine Fisheries 

Services, Seattle, Washington. 

Delfino, J.J. and G.F. Lee, 1971. Variation of manganese, 

dissolved oxygen and related chemical parameters in the bottom 

waters of Lake Mendota, Wisconsin. Water Res., 5:1207. 
DeMont, W.J., et al., 1975. Effect of atmospheric gas 

supersaturation caused by dams on salmon and steelhead trout to 

the Snake and Columbia Rivers. Final Report. Northwest 

Fisheries Center, NMFS, Seattle, Washington. 

De Sylva, D.P., 1969. Theoretical considerations of the effects 

of heated effluents on marine fisher. In: Biological aspects 

of thermal pollution. P.A. Krenkel and F.C. Parker (eds.), 

Vanderbilt University Press. 

photoperiod on reproductive cycling of the estuarine gobiid 

fish, Gillichthys mirabilk. Fishery Bull. 70, 4:1137. 

deVlaming, V.L., 1971. The effects of food deprivation and 

salinity changes on reproductive function in the estuarine 

gobiid fish, Gillichtys mirabilis. - Biol. Bull., 141:450. 

Diaz-Piferrer, 1962. The effects on an oil spill on the shore of 

Guanica, Puerto Rico (abstract) Ass. Isl. Mar. Labs, 4th Mtg. 

Curaco, 12-13. 

Doudoroff, P. and Katz, M.. 1953. Critical review of literature 

deVIaming, V.L., 1972. The effects of temperature and 

* 

- 

on the toxicity of industrial wastes and their components to 

fish. 11. The metals and salts. Sew. and Ind. Wastes, 25, 

p. 002. 

Doudoroff, P. and D.L. Shumway, 1970. Dissolved oxygen 

requirements of fresh water fishes. FA0 Fish. Tech. Paper 

No. 06. 

Dow, R.L., 1973. Fluctuations in marine species abundance dur 

climatic cycles. Mar. Tech. SOC. Jour., 7,4:30. 

Dowden, B.F., 1966. Effects of five insecticides on the oxygen 

consumption of the bluegill sunfish, Le omis macrochirus. 

Ph.D. Thesis, Louisiana State Univers fiton Rouge, LA. 

Eaton, J.G., 1970. Chronic malathion toxicity to the bluegill 

(Lepomis macrochirus Rafinesque) . Water Research, 4: 673. 

Ebel, W.J., et al., 1975. Effect of atmospheric gas 

supersaturation caused by dams on salmon and steelhead trout 

of the Snake and Columbia Rivers. Final Report. Northwest 

Fisheries Center, NMFS, Seattle, Washington. 

sediment studied in-situ and in the laboratory. Water 

Research, 7: 1285. 

Edberg, N. and Hofstan, B.V., 1973. Oxygen uptake of bottom 

Eddy, 

Univ., 

concentration 

chinook salmon 

R.M., 

June, 

1972. 

1972. 

and 

embryos 

The influence of 

temperature 

and fry. 

on the survival 

M.S. 

dissolved 

Thesis, Oregon State 

oxygen 

and growth of . 
Eichelberger, J.W. and J.J. Lichtenberg, 1971. Persistence of 

pesticides in river water. Environ. Sci. & Technol., 5:541. 

Eisler, R., 1969. Acute toxicities of insecticides to marine 

decapod crustaceans. Crustaceana, 16: 302. 

Eisler, ‘R., 1970. Acute toxicities of organochlorine and 

organophosphorus insecticides to estuarine fishes. U.S. 

Bureau of Sport Fisheries and Wildlife, Technical Paper 46. 

Ellis, M.M., 1937. Detection and measurement of stream 

pollution. Bull. US. Buraue of Sport Fisheries, and Wildlife. 

48 (22) : 365-437. 

Environmental Protection Agency, 1973. The control of pollution 

irom hydrographic modifications. EPA 430/9-73-017,. U.S. 

Government Printing Office, Washington, D.C. 

Environmental Protection Agency, 1985a. Technical Support 

Document for Water Quality-Based Toxics Control. Office of 

Water, Washington, D.C. 

Everett, G.V., 1973. Rainbow trout, Salmon gairdneri (Rich.) , 

fishery of Lake Titicaca. J. Fish. Biol., 5:429-440. 

European Inland Fisheries Advisory Commission, 1969. Water 

quality criteria for Eurpoean freshwater fish--extreme pH 

values and inland fisheries, prepared by EIFAC Working Party on 

Water Quality Criteria for European Freshwater Fish. Water 

Research, 3:593. 

ERL-Duluth, Environmental Reseach Laboratory, 1976, Procedures 

for developing temperature criteria for freshwater fish. 

Ecological Research Series, Report in Preparation. 

Fairbridge, R.W. (ed.) , 1966. The encyclopedia of oceanography. 

Reinhold, New York, New York. 

Federal Water Pollution Control Administration, 1967. 

Temperature and aquatic life. Laboratory Investigations-No.6, 

Technical Advisory and Investigations Branch, Cincinnati, Ohio. 

Fetterolf , C.M. Jr., 1973. Mixing zone concepts: Biological 

Methods for the Assesment of Water Quality, ASTM STP-528, 

American Society for Testing and Materials, pp. 31-45. 

Food and Drug Administration, 1974. Poisonous or deleterious 

substances in peanuts, evaporated milk, fish and shellfish. 

Proposed Rules. Federal Register, December 6, 1974, 

Washington , D.C. 

temperature, In: Thermobiology. A.H. Rose (ed.) , Academic 

Press, New York. 

Fry, F.E.J., 1967. Responses of vertebrate poikilotherms to 

Furch, X., 1972. Der Einfluss einer Vorbehandlung mit konstanten 
und wechselnden temperaturen auf die hitzerestistenz von 

Gammarus -salinus - und Idotea - balthica. Mar. Biol., 15:12. 

Gallagher, T.P., et al., 1969. Pollution affecting shellfish 

harvesting in Mobile Bay, Alabama, Tech. Servs. F.W.P.C.A. 

Southeast Water Lab., Athens, GA. 

Gammon, J.R., 1970. The effect of inorganic sediment on stream 

biota. Environmental Protection Agency. Water Poll. Cont. 

Res. Series, 18050 DWC 12/70, USGPO, Washington, D.C. 

Garside, E.T., 1966. Effect of oxygen in relation to temperature 

on the development of embryos of brook trout and rainbow trout. 

J. Fish. Res. Bd. Canada. 23:1121-1134. 

insecticides to various aquatic invertebrates. Water and Sew, 

Works, 112:276. 

Geldreich, E.E. and Kenner, 1969. Concepts of fecal streptococci 

in stream pollution. Jour. Water Poll. Contr. Fed. 41:R336. 

Geldreich, E.E., 1972. Buffalo Laboratory recreational water 

quality: A study in bacteriological data interpretation. 

Water Ree. 6:912. 

Geldreich, E.E., 1974. Microbiological criteria concepts for 

coastal bathing waters. Ocean Mgt. (In press). 

Geldreich, E.E. , 1974. Monitoring marine waters for 

microbiological quality. WHO Conference - Scientific Aspect of 

Marine Pollution. Geneva Switzerland. 

Gibson, J.R., - et --.I a1 1969. Sources of error in the use of fish- brain acetylcholinesterase activity as a monitor for pollution. 

Bull. Environ. Contam. Toxicol., 4:17. 

Gillette, L.A., et al.! 1952. Appraisal of a chemical waste 

problem by fish toxicity tests. Sewage Ind. Wastes, 24:1397. 

Glude, J.B., 1954. The effects of temperature and predators on 

the abundance of the soft-shell clam, arenaria, in New 

England. Trans. Amer. Fish. SOC., 04:13. 

Radioactivity in the marine enviroment. 

Sciences, Washington, D.C., p. 137. 

Gonzalez J.G., 1972. Seasonal variation in the responses of 

estuarine populations to heated water in the vicinity of a 

steam generating plant. Ph.D. Dissertation, Uinv. Rhode 

Island, p. 142. 

Gaufin, A.R., et al., 1965. The toxicity of ten organic 

Goldgerg, E.D., et al., 1971. Marine chemistry. In: 

National Academy of 

Griffin, A.E., 1960. Significance and removal of manganese in 
water supplies. Jour. Amer. Water Works Assn., 52:1326. 

Griffith, E.W., 1963. Salt as a possible limiting factor to the 

Suisan Marsh pheasant population. 

Wildlife Protection Study, Cooperative Study of California. 

Guarraia, L. J., 1972. Brief literature review of gebsiellg as 

pathogens. In seminar on the significance of Fecal coliform in 

Industrial Waste, E.P.A., T.R.3, National Field Investigations 

Center, Denver, CO, p. 94. 

- lucius L.) when incubated under different oxygen conditions. 

Probs. of Ichthyol., 9:841-851. 

Haas, A.R.C., 1932. Nurtirional aspects in mottleleaf and other 

physiological diseases of citrus. Hilgardia, 6:483. 

Hamdorf, K., 1961. Die Beeinflussung der Embryonal - und 

Larvalentqicklung der Regenbogenforelle (Salmo iridues Gibb.) 

durch die Umwelgfactoren 02-Partialdruck and Temperatur. 

verg 1. Physiol . 44 : 4 51-4 62. 

15:8. 

techniques for coagulation-filtration. jour. Amer. Water Works 

Assn., 59:1149. 

Harmeson, R.H., et al., 1971. The nitrate situation in Illinois. 

Jour. Amer. Water Works Assn., 63:303. 

Heath, R.G., et al., 1972. Comparative dietary toxicities of 

pesticides to birds. 

Wildlife Report No. 152, U.S. Dept. of the Interior, 

Washington, D.C., p. 51. 

environment; A continuing controversy. Intercont. Med. Book 

Corp., N.Y., pp. 421-435. 

concentrations on the growth of junenile coho salmon. Trans. 

Amer. Fish. SOC. 91:155-167. 

inorganic pollutants on young salmon and trout. 

Department Fish. Res., Bull. No. 5, p. 264. 

0 Annual Report Belta Fish and 

Gulidov, M.V., 1969. Embryonic develoopment of the pike (= 

2. 

Hampson, G.R., and H.L. Sanders, 1969. Local oil spill. Oceanus, 

Hannah, S.A., J.M. Cohen, and G.G. Robeok, 1967. Control 

Bureau of Sport Fisheries and Wildlife, 

Heath, R.G., and J.W. Spann, 1973. Pesticides in the 

Hermann, R.B., et al., 1962. Influence of dissolved oxygen 

Holland, G.A., et al., 1960. Toxic effects of organic and 

Washington, 

Holland, HeT.,, a&., 1967. Use of fish brain .L w acetylocholinesterase to monitor pollution by - organophosphorus 

pesticides. Bull. Environ. Contsm. Toxicol., 2:?~56.- 

HOSS, D.I., et al., 1975. Effects of temperature, copper and 
chlorine on fish during simulated entrainment in power plant 

condenser cooling systems. Symp. Physical and Biological 

Effects on the Environment of Cooling Systems and Thermal 

Discharges at Nuclear Power Plants. Intll Atomic Energy Agency 

(In press). 

comparison of total coliform and fecal coliform values in 

shellfish growing areas and a proposed fecal coliform growing 

area standard. Presented at 8th National Shellfihs Sanitation 

Workshop. (FDA, Washington, D.C.) 

Hunt, D.A. and J. Springer, 1974. Preliminary repdrt on tlA1t 

Hutchinson, G.E., 1957. A treatise on limnology. John Wiley & 

Hyde, X.M., et e., 1974. The effect of mirex on channel catfish 

Idler, D.R., 1969. Coexistence of a fishery and a major industry 

Illig, G.L. Jr., 1960. Significance and removal of manganese in 

Itazawa, Y. 1971. An estimation of the minimum level of 

Sons, New York. 

production, 

in Placentia Bay. Chemistry in Canada, 21(11):16. 

water supplies. Jour. Uer. Water works Assn., 52:1326. 

dissolved oxygen in water required for normal life of fish. 

Bull. Jap. SOC. Sci. Fish. 37:273-276. 

Trans. Amer. Fish SOC., 103:366. 

Jackim, E., et a1 ., 1970. Effects of metal poisoning for five 

1 iver enzymes in the kill if ish (Fundulus - heterocl itus). 

Fish. Res. Bd. Can., 27:383. 

Jour. 

Jacobson, S.M. and D.B. Boylan, 1973. Effect of seawater soluble 

fraction of kerosene on chemotaxis in a marine snail, Nassarius 

obsoletus. Nature, 241:213. 

Jangaard, P.M. 1970. The role played by the Fisheries Research 

Board of Canada in the Isredll herring phosphorus pollution 

crisis in Placentia Bay, Newfoundland. Fisheries Research 

Board, Atlantic Regional Office (Circular No. 1) Halifax, Nova 

Scotia. 

Jef fries, H.P. and W.C. Johnson, 1974. Seasonal distributions 

of bottom fishes in the Narragansett Bay area: Seven-year 

variations in the abundance of winter flounder 

(Pseudopleuronectes americanus) . Jour. Fish. Res. Bd. Can., 

31:1057. 

Jensen, L.D. and A.R. Gaufin, 1966. Acute and long-term effects 

of organic insecticides on two species of stonefly naiads. 

Jour. Water Poll. Cont. Fed., 38:1273. 
Johns Hopkins University, 1956. Final report to the water 

quality subcommittee of the American Petroleum Institute, 

Project, PG 49:41. 

Jones, J.R.E., 1964. Fish and river pollution. Butterworth 

and Co., Ltd., London. 

Kaiser, K.L.E., 1974. Mirex: an unrecognized contaminant of 

fishes from Lake Ontario. Science 185:523. 

Karnak, R.E. and W.J. Collins, 1974. The susceptibility to 

selected insecticides and acetylchlolinesterase activity in a 

0 

~ 

laboratory colony of midge larvae, Chironomus tentans 

(Dintera: chironomidae) . Bull. Enivorn. Contam. Toxicol., 

i2 :-a. 

Katz, M., 1961. Aucte toxicity of some organic insecticides to 

three species of salmonids and to the threespine strickleback. 

Trans. Amer. Fish. SOC., 90:264. 

Kelly, C.B. and W. Arcisy, 1954. Survival of enteric organisms 

in shellfish. Pub. Health Repts. 69~1205. 

Keup, L.E., 1975. Factors in fish kill investigations. Water 

and Sewage Works, 121:48. 

Kinne, O., 1970. Temperatur-animals-invertebrates, In: Marine 

ecology. 

Klinger, R., 1957. Sodium nitrite, a slow acting fish poison. 

Schweiz, 2. Hydrol. 19(2):565. 

Knepp, G.L. and G.F. Arkin, 1973. Ammonia toxicity levels and 

nitrate tolerance of channel catfish. The Progressive FishCulturist, 35: 221. 

insecticides to striped bass, Morone szxatilk. Calif. Fish 

and Game, 60:128. 

Kopfler, F.C., 1974. The accumulation of organic and inorganic 

mercury compounds by the eastern oyster (Crassostrea 

virginica). Bull. Environ. Contam. Toxicol. 11:275. 

United States. U.S. Dept. of the Interior, Federal Water 

Pollution Control Admin., Cincinnati, Ohio. 

Kovacs, N., 1959. Enteric fever in connection with pollution of 

seawater. Western Australia. Renort of the Commissioner Of 

0 0. Kinne (ed.), John Wiley and Sons, New York. 

Korn, S. and R. Earnest, 1974. Acute toxicity of twenty 

Kopp, J.F. and R.C. Kroner, 1967. Trace metals in waters of the 

klic Health for the year 1958. - 

0 
Kraus, as cited by Mitchess, J.W., R.E. Hogson, and C.R. 

Gaetjens, 1946. Tolerance of farm animals to feed containing 

2, 4-dichlorophenoxyacetic acid. Jour. Animal. Sci., 5:226. 

Lange, N.A., Ed., 1961. Handbook of chemistry, 10th ed. McGrawHill Book Co., New York. 

Langelier, W.F., 1936. The analytical control of anti-corrision 

water treatment. Jour. Amer. Water Works Assn., 28:1500. 

Le Clerc, E. and F. Devlaminck, 1955. Fish toxicity tests and 

water quality. Bull. de Belge Condumeiit Eaux., 28:ll. 

Le Clerc, E., 1960. The self-purification of streams and the 

relationship between chemical and biological tests. Proc. 

Second symposium on Treatment of Waste Waters, Pergamon Press, 

London, England, p. 281. 

Leggett, W.C. R.R. Whitney, 1972. Water temperature and the 

migrations of American shad, Fish. Bull. 70, 3:659. 

Lehman, A.J., 1965. Summaries of pesticide toxicity. 

Association of Food and Drug Officials of the U.S., Topeka, 

Kansas, pp. 1-40. 

pressure. Inst. Freshwater Res. Drottningholm, Report #29. 

Fish. Ed. of Sweden (Annual report for 1948), 49-50. 

water. Arch. Hydrobiology, 53 : 589. 

Lindroth, A., 1949. Vitality of salmon parr at low oxygen 

Lindroth, A., 1957. Abiogenic gas supersaturation of river 

Lockhart, E.E., - et -.I a1 1955. The effect of water impurities on 

Lowe, J.I., et al., 1971. Effects of mirex on selected estuarine 

the flavor of brewed coffee. Food Research, 20:598. 

organisms. In: Transactions of the 36th North American 

Wildlife Resources Conference. pp. 171-186. 

Lowman, F.G.,, et al., 1971. Accumulation and redistribution of 

radionuclides by marine organisms, In: Radioactivity in the 

Marine Environment, National Academy of Sciences, Washington, 

D.C., p. 161. 

Ludke, J.L., et al., 1971. Toxicity of mirex to crayfish, 

Procambarus blandingi. - Bull. Environ. Contam. Toxicol., 6:89. 

Lurid, W.A. and G.C. Maltezos, 1970. Movements and migrations of 

the bluefish, Pomatomus etarix, tagged in waters of New York 

and southern New England. Trans. Amer. Fish. SOC., 99:719. 

Macek, X.J. and McAllister, 1970. Insecticide susceptibility of 

some common fish family representatives. Trans. her. Fish 

SOC., 99:20. 
Macek, K.J., et al., 1974. Chronic toxicity of lindane to 

selected aquatic invertebrates and fishes. Environmental 

Protection Agency Research Contract Report, Enviornmental 

Protection Agency Ecol. Res. Series, (In preparation). 

Mackenthun, K.M., 1973. Toward a cleaner aquatic environment. 

US. Government Printing Off ice, Washington, D.C. 

Malouf, R., 1972. Occurrence of gas bubble disease in three 

species of bivalve mollusks. Jour. Fish. Res. Bd. Can., 

0 

29:588. 

Marcello, R.A., et al., 1975. Evaluation of alternative 

solutions to gas buble disease mortality of menhaden at Pilgrim 

Nuclear Power Station. Yankee Atomic Electric Co., Westboro, 

Mass. YAEC-1087. 

Massengill, R.R., 1973. Change in feeding and body condition of 

brown bullheads overwintering in the heated effluent of a power 

plant. Ches. Sci. 14,2:138. 

McCord, C.P., 1951. Beryllium as a sensitizing agent. Ind. Med. 

Surg., 20:236. 

McCoy, E.F., 1972. Role of bacteria in thenitrogen cycle in 

lakes. Enviornmental Protection Agency, Water Pollution 

Control Research Series, U.S. Govenrment Printing Off ice (EP 

2.10:16010 EHR 03/72). Washington, D.C. 

McKim, J.M., et al., 1976. Long-term effects of methylmercuric chloride on three generations of brook trout 

(Salvelinus __ fontinalis) : Toxicity, accumulation, 

distribution, and elsination. Jour. Fish. Res. Board Can. 

33:2726. 

McKee, J.E. and H.W. Wolf, 1963. Water quality criteria. State 

Water Quality Control Board, Sacramento, CAI pub. 3-A. 

Merna, J.W. and P.J. Eisele, 1973. The effects of methoxychlor 

on aquatic biota. U.S. EPA Ecological Res. Series, No. EPA-R3- 73-046. U.S. Government Printing Office, Washington, D.C. 

Bull. World Health Organization, 44:363. 

shellfish in esturary waters. Jour. San. Eng. Div. Proc. Amer. 

SOC. Civ. Eng., 94:595. 

selective fish eradicator. Trans. Amer. Fish. SOC., 94:203. 

Metcalf, R.L., & a&., 1971. Biodegradable analogs of DDT. 

Metcalf, T.G. and W.C. Stiles, 1968. Viral pollution of 

Meyer, F.P., 1965. The experimental use of guthion as a 
Minchew, G.D. and D.E. Ferguson, 1970. Toxicities of six 

insecticides to resistant and susceptible green sunfish and 

golden shiners in static bioassays. Jour. Miss. Acad. Sci., 

15:29. 

Mironov, O.G., 1967. Effects of low concentrations of petroleum 

and its products on thedevelopment of roeoftheBlackSea 

flatfish. Vop Ikhtiol., 7:557. 

Mironov, O.G., 1970. The effect of oil pollution on flora and 

fauna of the Black Sea. In Poceedings: FA0 Conference on 

Marine Pollution and its effects on living resources and fish. 

Rome, Dec., 1970., E-92. Food and Agriculture Organization of 

the United Nations. 

Moore, E.W., 1952. Phisiological effects of the consumption of 

saline drinking water. National Res. Council, Div. of Medical 

Sciences, Bull. San. Engr., and Environment, Appendix E. 

Moore, S.F., et al., 1973. A preliminary assessment of the 

environmental vulnerability of Machias Bay, Maine to oil super- tankers. 

Mount, D.I., 1960. Effects of various dissolved oxygen levels on 

fish activity. Ohio State Univ. Natur. Resources Inst., Ann. 

Fisheries Res. Rept. pp. 13-33. 

freshwater fishes. In: Biological aspects of thermal 

pollution. P.A. Krenkel and F.L. Parkers, (eds.) , Vanderbilt 

University Press. 

survival, growth and reproduction. Water Res., 7:987. 

acceptable toxicant limits for fish-malathion and the 

butoxyethanol ester of 2,4-D. Trans. Amer. Fish. SOC., 

21:185. 

Mulla, M.S. and A.M. Khasawina, 1969. Laboratory and field 

evaluations of larvicides against chironomid midges. Jour. 

Econ. Entomol., 62:37. 

National Academy of Sciences - Conmittee on Biologic Effects of 

Atmospheric Pollutants, 1972. Lead: Airborne lead in 

perspective. The National Academy of Sciences, Washington, 

D.C. 

National Acadmey of Sciences, National Academy of Engineering, 

Mount, D.I., 1969. Developing thermal requirements for 

Mount, D.I., 1973. Chronic effect of low pH on fathead minnow 

Mount, D.T. and C.E. Stephan, 1967. A method for establishing 

1974. Water quality criteria, 1972, U.S. Government Printing 

Off ice, Washington, D.C. 
National Institute for Occupational Safety and Health, 1972. 

Occupational exposure to beryllium. U.S. Department of Health, 

Education and Welfare, Health Services and Mental Health 

Admin., National Institute for Occupational Safety and Health. 

U.S. Government Printing Off ice, Washington, D.C. 

National Research Council, 1954. Sodium restricted diets. 

Publication 325, Food and Nutrition Board, Washington, D.C. 

National Technical Advisory Committee to the Secretary of the 

Interior, 1968. Water quality criteria, U.S. Government 

Printing Office, Washington, D.C. 

Nagvi, S.M. and D.E. Ferguson, 1969. Pesticide tolerances of 

selected freshwater invertebrates. Jour. Miss. Acad. Sci. , 

14 : 121. 

Nebeker, A.V. and A.R. Gaufin, 1964. Bioassays to determine 

pesticide toxicity to the amphipod crustacean, Gammarus 

lactuis. Proc. Utah Acad. Sci. Arts and Letters, 41:64. 

Nebeker, A.V., et al., 1976a. Nitrogen, oxygen, and carbon 

dioxide as factors affecting fish survival in gas 

supersaturated water. Trans. Amer. Fish. SOC. (In Press). 

salmon molts in seawater after exposure to gas supersaturated 

water. Trans. Amer. Fish. SOC. (In Press). 

water on freshwater invertebrates. Proc. Gas Bubble Disease 

Workshop. Battelle Northwest, ERDA Special Report (In Press). 

Neill, W.H. and J.J. Magnuson, 1974. Distributional ecology and 

behavioral thermoregulation of fishes in relation to heated 

effluent from a power plant at Lake Monona, Wisconsin. Trans. 

Amer. Fish. SOC. 103, 4:663. 

animals, p. 273-380. In: P. Hepple (ed.), Water Pollution by 

oil. I.P. London. 

Copiea No., 55:37. 

chloramphenical, amphicil lin and other anrimicrobial agents: 

Strains isolated during an extensive typhoid fever epidemic in 

Mexico. Antimicrobial Agents and Chmeothearpy, 4:597. 

0 

Nebeker, A.V., et al., 197633. Survival of coho and sockeye 

Nebeker, A.V., et al., 1975. Effects of gas supersaturated 

Nelson-Smith, A., 1971. Effects of oil on marine plants and 

Nichols, J.T., 1918. An abnormal winter flounder and others. 

Olearte, J. and E. Galindo, 1973. - Salmonella typhi resistance to 

Olla, B.L. and A.L. Studholme, 1971. The effect of temperature 

on the activity of bluefish, Pomatomus saltatrix (1) Biol. 

Bull., 141:337. 
0 Olson, G.F., -_ et al., 1975. Mercury residues in fathead minnows, 

Pimephales promelas Rafinesque, chronically exposed to methyl- mercury in water. Bull. Environ. Contam. Toxcol., 14:129. 

hydrogen sulfide to Gammarus psedol - imnaeus. Trans. her. Fish. 

sac., 103 (In press). 

Oseid, D.M. and L.L. Smith, Jr., 197a. Chronic toxicity of 

Oseid, D.M. and L.L. Smith, 197413. Long-term effects of hydrogen 

sulfide on Hexaqenia limbata (Ephemeroptera) . Environmental 

Ecology (In press). 

oseid, D.M. and L.L. Smith, 1974c. Factors influencing acute 

toxicity estimates of hydrogen sulfide to freshwater 

invertebrates. Water Research 8 (In press). 

Palmer, J.S. and R.D. Radeleff, 1964. The toxicologic effects of 

certain fungicides and herbicides on sheep and cattle. Ann. 

N.Y. Acad. Sci., 111:729. 

Paris, D.F., et all 1975. Rates of degradation of malathion 

by bacteria isolated from aquatic systems. Environ. Sci. & 

Technol., 9:135. 

Parker, F.L. and P.A. Krenkel, 1969. Thermal pollution: Status 

of the art. Report No. 3, Vanderbilt University, School of 

Engineering, Nashville, Tennessee. 

Patrick R., 1969a. Some effects of temperature on freshwater 

algae. In Biological aspects of thermal pollution. P.A. 

Krenkel and F.L. Parker, (eds.) , Vanderbilt University Press. 

Patrick, R., 1969b. Discussion of engineering aspects, soucres 

and magnitude of thermal pollution. By P.A. Krenkel and F.L. 

Parker. In: Biological aspects of thermal pollution. P.A. 

Krenkel and F.L. Parker, (eds.) , Vanderbilt University Press. 

mineralized water on houshold plumbing and appliances. Jour. 

Amer. Water Works Assn., 60:1060. 

Pearce, J.B., 1969. Thermal addition and the benthos, Capd Cod 

Canal. Ches. Sc., 10:227. 

Pearse, J.S., 1970. Reproductive periodicities of Indo-Pacif ic 

invertebrates in the 6ulf of Suez. 111. The echinoid Diadema 

setosum (Leske). Bull. Mar. Sci., 20, 3:697. 

organisms after contamination of the bottom of Long Harbour, 

Placentia Bay, Newfoundland with elemental phosphorus. In: 

Effects of Elemental Phosphorus on Marine Life, Fish. Res. Bd4 

of Canada, Circular 2, pp. 181-186. 

Patterson, W.L. and R.F. Banker, 1968. Effects of highly 

Peer, D.L., 1972. Observations on mortalities of e\benthis 
Phelps, E.B., 1944. Stream sanitation. John Wiley and sons, 

Phillips, A.M., Jr., D.L. Livingston and R.F. Dumas, 1960. 

Inc. , New York. 

Effects of starvation on the chemical composition of brook 

trout. Prog. Fish. Cult., 22:147. 

phosphorus inseFticides to different species of warmwater 

fishes. Trans. Amer. Fish. SOC., 91:175. 

Plotkin, S.A. and M. Katz, 1967. Minimal infective doses of 

viruses for man by the oral route. In: Transmission of 

Viruses by the Water Route, G. Berg, Ed., John Wiley 

Interscience, N.Y., p. 155. 

using standard methods. In: Marine pollution and sea life 

(Ruivo, Ed.). Fishing News (Books) Ltd, London, England, 

pp. 212-217. 

Pickering, Q.H., et a&., 1962. The toxicity of organic 

Portman, J., 1972. Results of acute tests with marine organisms, 

Post, G. and T.R. Schroeder, 1971. The toxicity of four 

insecticides to four salmonid species. Bull. Environ. Contam. 

Toxicol., 6:144. 

Post, G. and R.A. Leasure, 1974. Sublethal effect of malathion 

to three salmonid species. Bull. Environ. Contam. Toxicol., 

12:312. 

Presnell, N., 1974. Discussion of fecal coliforms for shellfish 

growing waters. Proc. 7th National Shellfish Sanitation 

Workshop, Oct. 21-22, 1971. Ratcliff and Wilt, Eds., FDA., 

Washington, D.C. 

Raney, E.C., 1969. Discussion of effects of heated discharges 

on freshwater fish in Britian. By J.S. Alabaster. In: 

Biological aspects of thermal pollution. P.A. KrenkeL and 

F.L. Parker (eds.) , Vanderbilt University Press. 

Rawson, D.S. and J.E. Moore, 1944. The saline lakes of 

Saskatchewan. Canadian Jour. of Res. , 22 : 141. 

Reeves, A.L., 1965. Absorption of beryllium from the 

gastrointestinal tract. A.M.A. Arch. Envir. Health, 11:209. 

Reid, L.C. and D.A. Carlson, 1974. Chlorine disinfection of low 

temperature waters. Jour. Environ. Eng. Div., ASCE, Vol. 100, 

No. EE2:339. 

Reitler, R. and R. Seligaman, 1957. Pseudomonas aeruginosa in 

drinking water. Jour. Appl. Bact., 20:145. 

Renfro, W.C., 1963. Gas bubble mortality of fishes in Galveston 

Bay, Texas. Trans. Amer. Fish. SOC., 92:320. 
Richards, L.A. (Ed.) , 1954. Diagnosis and improvement of saline ., 

and alkali soils. Agriculture Handbook. No. 60, U.S. Government 

Printing Office, Washington, D.C. 

humans. Am. J. Physiol., 126:l. 

supplies. Jour. Amer. Water Works Assn., 50:680. 

elements in species of phytoplankton grown in culture. Jour. 

Mar. Biol. Assn. U.X., 51:63. 

Ringem, L.M. and C.H. Drake, 1952. A study of the incidence of 

Ricter, C.O., and A. MacLean, 1939. Salt taste threshold of 

Riddick, J.M., et al., 1958. Iron and Manganese in water 

Riley, J.P. and I. Roth, 1971. The distribution of trace 

Pkeudomonas aeruginesa from various national sources. Jour. 

Bact. , 64 :841. 

Romney, E.M., et al., 1962. Beryllium and the growth of bush 

Romney, E.M. and J.D. Childress, 1965. Effects of beryllium in 

beans. Science, 135:786. 

'plants and soil. Soil Sci., 100:210. 

Rounsefell, G.A. and W.H. Everhart, 1953. Fishery science, its 

methods and applications. John Wiley and Sons, Inc., New York. 

Rucker, R.R., 1974. Gas bubble disease: Mortalities of coho 

salmon, Oncorhynchus kisutch, in water with constant total 

gas pressure and different oxygen-nitrogen ratios. National 

Oceanic and Atmas. Admain., Natl. Mar. Fish. Serv., Northwest 

Fish Center, Seattle, Washington, unpublished manuscript. 

0 

Rulifson, R.L. and G. -el, 1971. Nitrogen suspesaturation in 

the Columbia and Snake Rivers. Tech Rept. TS-09-70-208-016.2, 

Environmental Protection Agency, Region X, Seattle, Washington. 

trout (Salmon ---- gaikdneri). --- 

Russo, R.C., et al., 1974. Acute toxicity of nitrite to rainbow 

Jour. Fish. Res, Bd. Can., 31:1653. 

Russo, R.C. and R.V. Thurfiton, 1975, Acute toxicity of nitrite 

to cutthroat trout (Salmon clarki). Fisheries Bioassay 

Laboratory Tech. Report No. 75-3, Montana State University. 

Saeki, A., 1965. Stu6ie.s on fish culture in filtered closedcirculating aquaria. 11. on the carp culture experiments in 

the systems. Bull. Jap. Soa. Sci. Fish., 31:916. 

Sanders, H.O., 1969. Toxicity of pesticides to the crustacean, 

Gammarus lacustris. US. Department of the Interior, 

Washington, D.C., Bureau of Sport Fisheries and Wildlife 

Technical Paper -No. 25. - 
Sanders, H.O., 1972. Toxicity of some insecticides to four - species of malacostracan crustaceans. U.S. Department of the 

Interion, Washington, D.C. Bureau of Sport Fisheries and Wild- life Technical Paper No. 66. 

pesticides to two species of cladocerans. Trans. Amer. Fish. 

Sanders, H.O. and O.B. Cope, 1966. Toxicities of several ' 

SOC., 95:165. .? 

Sanders, H.O. and O.B. Cope, 1968. The relative toxicities of 

several pesticides to naiads of three species of stonefly. 

Limnol. & Oceanong., 13:112. 

Sastry, A.N., 1975. Physiology and ecology of reproduction in 

marine invertebrates. In: F.J. and W.B. Vernberg (eds.) , 

Physiological ecology of estuarine organisms. Belle W. Baruch 

Library of Marine Species, Univ. South Carolina Press, 

Columbia, South Carolina. 

Sattlemacher, P.G., 1962. Methemoblobinemia from nitrates 

in drinking water. SchrReihe. Ver. Wasser-, Boden-u. Luthyg. 

No. 21, Fishcer, Stuttgart. 

Hi1 1 , New York. 

as indicators of sewage pollution. 

from Sear Outfalls. Pergamon Press. 

Seabury, J.H., 1963. Toxicity of 2, 4-dichlorophenoxyacatic acid 

for man and dog. Arch. Envir. Health, 7:202. 

Shumway, D.L. and J.R. Palensky, 1973. Impairment of the flavor 

of fish by water pollutants. U.S. Environmental Protection 

Agency, EPA-R3-73-OlOI U.S. Government Printing Office, 

Washington, D.C. 

Shuval, H.J., et aJ., 1971. Natural inactivation processes 

of viruses =seawater. Proc. Natural Specialty Conf. on 

Disinfection. Amer. SOC. of Civil Engineers, N.Y. 

Saywer, C.N., 1960. Chemistry for sanitary engineers. McGrawScarpino, D., 1974. Human enteric virsuses and bacteriolphages 0 International Symposium 

Siefert, R.E., et al., 1973. Effects of reduced oxygen 

concentrations on northern pike (Esox -lucius) embryos - and 

larvae. J. Fish. Res. Bd. Canada. 30:849-852. 

concentrations on the early life stages of mountain whitefish, 

smallmouth bass, and white bass. Accepted for publication in 

the Prog. Fish. Cult. 

Siefert, R.E., et al., 1975. Effects of reduced oxygen 
Siefert, .RE. and W.A. Spoor., 1974. Effects of reduced oxygeriie 

concantrations on embryos and larvae of white. sucker, coho 

salmon, brook trout, and walleye. Procaedings'..of: an ' 

International Symposium on the Early Life History of Fish. 

Springer-Verlag Berlin Heidelberg New York. 

diseases sf herring. from the western North Atlantic.. Spec.: 

Publ. Corn. Horthw. Atlantic Fish. No. 6:603-610. 

Slanetz, L.W., et a1 1965. Correlation of coliform and fecal 

streptococci indices .with 'the presence of Salmonella -- and, , ' 

enteric viruses in sea water and she2lfish. .Adv....in Water 

Pollution Res. 2nd International Confr., Tokyo, 3:17. 

Slonim, A.R:, 1973. Acute toxicity of beryllium sulfate to the 

common guppy. Jour. Water Poll. Cont. Fed. 45:2110. 

Slonim, A.R. and E.E. Ray, 1975. Acute.toxic y of beryllium 

sulfate to salamander larvae (Ambystoma q~). ' ' Bull. Envir. 

Contam. Toxicol. , 13 : 307. 

Slonim, C.B. and A.R..; sibnim, 1973.. Effecf':oi 5 hardness ,on 

the tolerance of the guppy to berylliumzsul~f Bull. Envir. 

Contam. Toxicol., 10:295. 

Smith, R.S., et al., 1951. Bathing water. quality and he.alth. 

I. Great Lakes (U.S. Public Health Service, Cincinnati, ohio). 

Smith, L.L., 1971. Influence of hydrdgen. sulfide .bn'. fish and -' 

arthropods. Environinental Protection Agency, :, Project 18050 

PCG, Washington, D.C. 

Smith, L.L. and' D.K. Oseid, 1972. Effects of .hydrogen sulfide-& 

fish eggs and fry. Water Research, 6:7P1. 

Smith, R.S. and T.B. Woolsey, ,1932. ''"'Bathing wator quality and.'. 

health. 11. Inland river and pool. '(U.S.. ,&bl'ic Health I 

Smith, K.S. and T.D. Woolsey, 1961. Bathing 'water quality and:" 

'Obarr, Scotland, May 17-23, 1973, Edited by J.H.S. .,Blaxter, '* <. 

pp.- 487-495, .. , 

Sinderman, C.J., 1965. Effects of environment on several 

: .. .. 

- - -,.? -.I 

. .. , . ... . ! > " r;, ,, 

.. 

.. 

., 

,. .. Service, Cincinnati, Ohio). ~ . ., 

Public Health. III. Coastal Waters (U.5; 'Public Health 

Snow, J.R., 1958. A preliminary report on the comparative 

.-, .. .. .. .. .. ., .. . 1. 

.., .. . Serviee, Cincinnati, Ohio). .. 

testing of.some of the newer herbicides. Proc. 11th Ann. 

.'Conf., Southeast ASSOC. Game Fish COI~IIB.,. pp. ,125-132. 

, Report of the Commissioner 'uf hblic Healt i .... 

.. 

Snow, D.J.R., 1959. Typhoid and City Beach. Western Australia, 

or the year 1958; 

p. 52. 

Phil adel phia, Pennsylvania. 

Sollman, T.H.., 1957. A manual of pharmacology. W.B. Saunders, 
South, G.E. and R.D. Hill, 1970. Stud$e,s. on marine a.lgae.of.. -: 

Newfoundland. I. Occurrence and -distr.ibution of-, free .... ..iiG.hg ., .' 

Ascophyl __ lum nOdOSum in Newfoundian&" ... : Can. Jour. Bot., .............. ./ 46:1697. 

Sprague, J.B., 1963. Resistance .'& f6ur:.fresh, wat 

to lethal, high temperature and, low ... oxygen. J.. 

Canada. , 20: 387. 

1971. 13th Edition, APHA, Amer. .- ... ; 

Cont. Fed. 

.'~ .... 

.. 

Standard. Methods for the Examinatio 

- 

Stev~enson, A.H., - 1.953. ~ Studies ..of bathing water. quality ... and ' '' 

Stewart, B.A., et al., i96 N'itrate and other p utants under 

fields and feedl.ots.. .~ . .Emir.. _. Sci. .... Tech.,:, ... .1:73.6.. ... 

Stewart, N.E., et' al., 1967. Influence of oxygen concentration 

on the growth of juvenile.,.largemouth bass,, J. Fish., Re's. Bd. 

Canada., 24~475-494. 

.. health.. Amer, Sour. 'hblic .- .... Health, 43~529; ........ ... .. 

- .., ,.. . ......... ........ . . 

.. .. ........ ~ . 

. ,. , 

Stickel, L.F., 1973. Pesticide residues in- bi'rds.and mammals. ~ In: Environmental .,po&lution ides.,Jc.p. Edwards, Ed..) 

Stiff, M.J., 1971. Copper/bicarbonate equilibria in solutions of 

Plenum Pxess, New York. pp., ..... :. . .. .... 

bicarbonate ion at concentrations simi.lar t,o tho.se found in 

natural water. . Water ... Research, ... ........ 5:171.: ...... 

experimental beryl 1i.m poisoning. Jour.: ...... Lab. C1 in. Med., 

38: 173. 

and sons, Inc., New ... york', ,.Chapter'4. .. 

I iiiting factor.; r. striped . .., ..... bass. . ' Tran 'her. Fish. soc., 

Special Pub., 3:3i: ... .. ..... 

Stokinger, H.E. and C.A.:.Stroud, 1951, Anemia in acute 

... 

.Stiim, W. and J.J. Morgan, .L970. Aquat.ic-. ...... Chemistry.. John Wiley 

Talbot, G.B., ,1966 ,..,. stuarine environmental requirements and . -. 

. .. 

. I .. 

Tarzwell, C.M. .and A.R.. Gaufin, . ,1953.. some^ important biolbgica3. ..- 

effects of pollution .otte 

proceedings of the 8th PU 

Reprinted in Biology of Water Pollution, i967. DepC Of 

Interior, Washington, - D.c. . .. , .. .. 

sregarded in stream surveys, ., %.: 

indistriai .. 'waste confeyence.,, .. ~ 

. .- , ... . J. '2 

Tarzwell, C.M. and C. Henderson,. 1960. 

$aylor, E.W., 1968. $oFdrty-secop f . Director of Water:. .-' 

Toxicity .. of less'.co . .~-. . , metals to fishes. ' Industrial .Wastes, 5i12. 

Examination, 19 65-66 , Metropo er Board, Londori,'~$:~ 117. 

.. ....... ..... ...... .. . 

. -. _. ^, *.> 

.... .... .. I. ... .~ ?.~,.:-. 

.. .. =.Vc - .: .. 

> 

. ,- .. . ._. 
U.S. Bureau of Sp.o.rt Fisheries and Wildlife, 1970. Resource 

wa&hinGtok, &Q~;. .?!<.I . i: . Ci 

..... .. .... ii_ _:_.. :.-i:2 ::i ?;,,-.:,:,~.v:%.:; 2,- 

U.S. Environmental Protection Agency. 1985s. Techiii&alF*Sup@oft 

Document for Water Qualitv-Based Toxics Control. Office of 

.. 5 ,,'.2<, ... \_"> L.. : i, ...... :__- I 
Van Donsel, .D:.J..- :and. .E.E. . Geld Relations.bips of c,dciT; 

,to gecal. colifo sediments,: .. - W$*k- .BgS:'- . . . &',. -1 v; .- ,. . ... .~.2 ..,'.,'~IZ si; .. .-~ 

Van Horn, W.M., 1958. 

Van Ords+%nd, 'H. 

The effect of pul~p and paper mill wastes" 

on aquatic.- 1.iif.e . ,.€'roc. .. Ontari nf.,, 5~50. ,. . s..;:<ie.ri: .. :.- .. 

Amer. Med. Assn., 1=:1084. 

van &?a 1,in ,.. C,zF. 

w.ate.r .f iqhes. 

. 

a0:1119. 

nput_qutput, mqdeJ!s.r Schweiz 2. 

waters. Sewage Ind. Wastes, 

: .r+~.at+ng ..,to infant 

aminated water. Amer.""' 

..' i.'( ,~,. , . . .~ . ~*.Lli ::. 

2 9.:4 9 5 . .. ,.. 

. . 

, ~ .. . ., . . _._ 

- I.. ., :",L,-., .,.* ~ - - r.." ...- ~. .. . i, 

*,uJ. QOVERNMENT PRlNTlNQ OFflCE] 9 8 6 -1 5 9- 3 0 0 5 0 4 7 2 
QUALITY CRITERIA FOR WATER 1986 

UPDATE #I 
*ALDRIN-DIELDRIN 

CRITERIA: 

Aquatic Life 

Dieldrin 

For dieldrin the criterion to protect freshwater aquatic life 

as derived using the Guidelines is 0.0019 ug/L as a 24-hour 

average, and the concentration should not exceed 2.5 ug/L at any 

time. 

For dieldrin the criterion to protect saltwater aquatic life 

as derived using the Guidelines is 0.0019 ug/L as a 24-hour 

average, and the concentration should not exceed 0.71 ug/L at any 

time. 

Aldrin 

For freshwater aquatic life the conckntration of aldrin 

should not exceed 3.0 ug/L at any time. No data are available 

concerning the chronic toxicity of aldrin to sensitive 

freshwater aquatic life. 

For saltwater aquatic life the concentration of aldrin should 

not exceed 1.3 ug/L at any time. No data are available 

concerning the chronic toxicity of aldrin to sensitive saltwater 

aquatic life. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to aldrin through ingestion of 

contaminated water and contaminated aquatic organisms, the 

*Indicates suspended, canceled or restricted by U.S. EPA office 

of Pesticides and Toxic Substances 
ambient water concentration should be zero, based on the 

nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, 

the levels which may result in incremental increase'of cancer 

risk over the lifetime are estimated at and 

The corresponding recommended criteria are 0.74 ng/L, 0.074 ng/L, 

and 0.0074 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption 

of water, the levels are 0.79 ng/L, 0.079 ng/L, and 0.0079 

ng/L, respectively. 

0 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to dieldrin through ingestion of 

contaminated water and contaminated aquatic organisms, the 

ambient water conceqtration should be zero, based on the 

nonthreshold assumption for this chemical. However, zero level 

may not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

lifetime are estimated at loe6 and The 

corresponding recommended criteria are 0.71 ng/L, 0.071 ng/L, and 

0.0071 nq/L, respectively. If these above estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 0.76 ng/L, 0.076 ng/L, and 0.0076 ng/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
AMMONIA 

sUMlrmy: 

~11 concentrations used herein are expressed as un-ionized 

ammonia (NH3), because NH3, not the ammonium ion (NH4') has been 

demonstrated to be the principal toxic form of ammonia. The 

data used in deriving criteria are predominantly from flow 

through tests in which ammonia concentrations were measured. 

Ammonia was reported to be acutely toxic to freshwater organisms 

at concentrations (uncorrected for pH) ranging from 0.53 to 22.8 

mg/L NHj for 19 invertebrate species representing 14 families and 

16 genera and from 0.083 to 4.60 mg/L NH3 for 29 fish species 

from 9 families and 18 genera. Among fish species, reported 96- 

hour LC50 ranged from 0.083 to 1.09 mg/L for salmonids and from 

0 

0.14 to 4.60 mg/L NH3 for nonsalmonids. Reported data from 

chronic tests on ammonia with two freshwater invertebrate 

species, both daphnids, showed effects at concentrations 

(uncorrected for pH) ranging from 0.304 to 1.2 mg/L NH3, and 

with nine freshwater fish species, from five families and seven 

genera, ranging from 0.0017 to 0.612 mg/L NH3. 

Concentrations of ammonia acutely toxic to fishes may cause 

loss of equilibrium, hyperexcitability, increased breathing, 

cardiac output and oxygen uptake, and, in extreme cases, 

convulsions, coma, and death. At lower concentrations ammonia 

has many effects on fishes, including a reduction in hatching 

success, reduction in growth rate and morphological development, 

and pathologic changes in tissues of gills, livers, and kidneys. 
Several factors have been shown to modify acute NH3 toxicity 

in fresh water. Some factors alter the concentration of unionized ammonia in the water by affecting the aqueous ammonia 

equilibrium, and some factors affect the toxicity Of un-ionized 

ammonia itself, either ameliorating or exacerbating the effects 

of ammonia. Factors that have been shown to affect ammonia 

toxicity include dissolved oxygen concentration, temperature, 

p~, previous acclimation to ammonia, fluctuating or intermittent 

exposures, carbon dioxide concentration, salinity, and the 

presence of other toxicants. 

The most well-studied of these is pH; the acute toxicity of 

NH3 has been shown to increase as pH decreases. Sufficient data 

exist from toxicity tests conducted at different pH values to 

formulate a mathematical expression to describe pH-dependent 

acute NH3 toxicity. The very limited amount of data regarding 

effects of pH on chronic NH3 toxicity also indicates increasing 

NH3 toxicity with decreasing pH, but the data are insufficient 

to derive a broadly applicable toxicity/pH relationship. Data on 

temperature effects on acute NH3 toxicity are limited and 

somewhat variable, but indications are that NH3 toxicity to fish 

is greater as temperature decreases. There is no information 

available regarding temperature effects on chronic NH3 toxicity. 

Examination of pH and temperature-corrected acute NH3 

toxicity values among species and genera of freshwater organisms 

showed that invertebrates are generally more tolerant than 

fishes, a notable exception being the fingernail clam. There is 

no clear trend among groups of fish; the several most sensitive 
tested species and genera include representatives from diverse 

faailies (Salmonidae, Cyprinidae, Percidae, and Centrarchidae). 

Available chronic toxicity data for freshwater organisms also 

indicate invertebrates (cladocerans, one insect species) to be 

more tolerant than fishes, again with the exception of the 

fingernail clam. When corrected for the presumed effects of 

temperature and pH, there is alsonoclear trend among groups of 

fish for chronic toxicity values, the most sensitive species 

including representatives from five families (Salmonidae, 

Cyprinidae, Ictaluridae, Centrarchidae, and Catostomidae) and 

having chronic values ranging by not much more than a factor or 

two. The range of acute-chronic ratios for 10 species from 6 

families was 3 to 43, and acute-chronic ratios were higher for 

the species having chronic tolerance below the median. 

Available data indicate that differences in sensitivities between 

warm and coldwater families of aquatic organisms are inadequate 

to warrant discrimination in the national ammonia criterion 

between bodies of water with "warm" and "coldwater" fishes: 

rather, effects of organism sensitivities on the criterion are 

most appropriately handled by site-specific criteria derivation 

procedures. 

Data for concentrations of NH3 toxic to freshwater 

phytoplankton and vascular plants, although limited, indicate 

that freshwater plant species are appreciably more tolerant to 

NH3 than are invertebrates or fishes. The ammonia criterion 

appropriate for the protection of aquatic animals will therefore 

in all likelihood be sufficiently protective of plant life. 
Available acute and chronic data for ammonia with saltwater 

organisms are very limited, and insufficient to derive a 

saltwater criterion. A few saltwater invertebrate species have 

been tested, and the prawn Macrobrachiurn _rosenbergiA _-- was the 

most sensitive. The few saltwater fishes tested suggest greater 

sensitivity than freshwater fishes. Acute toxicity of NH3 

appears to be greater at low pH values, similar to findings in 

freshwater. Data for saltwater plant species are limited to 

diatoms, which appear to be more sensitive than the saltwater 

invertebrates for which data are available. 

More quantitative information needs to be published on the 

toxicity of ammonia to aquatic life. Several key research needs 

must be addressed to provide a more complete assessment of 

ammonia toxicity. These are: (1) acute tests with additional 

saltwater fish species and saltwater invertebrate species; (2) 

life-cycle and early life-stage tests with representative 

freshwater and saltwater organisms from different families, with 

particular attention to trends of acute-chronic ratios; (3) 

fluctuating and intermittent exposure tests with a variety of 

species and exposure patterns; (4) more complete tests of the 

individual and combined effects of pH and temperature, especially 

for chronic toxicity; (5) more histopathological an6 

histochenical research with fishes, which would provide a rapid 

means of identifying and quantifying sublethal ammonia effects; 

and (6) studies on effects of dissolved and suspended solids on 

acute and chronic toxicity. 
NATIONAL CRITERIA: 

The procedures described in the Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses indicate that, except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if: 

8 

(1) the 1-hour* average concentration of un-ionized ammonia 

(in mg/L NH3) does not exceed, more often than once every 3 years 

on the average, the numerical value given by 0.52/FT/FPH/2, 

where: 

FT - - 10°.03 20-TCAP); TCAP < T 5 30 - 

10 0.03 20-T); 0 < T < TCAP - - 

FPH = 1 ;8<pH<9 

1+107.4-PH 

1.25 ; 6.5<pH - - < 7.7 

TCAP = 20 C; Salmonids or other sensitive 

coldwater species present 

coldwater species absent = 25 C; Salmonids and other sensitive 

(*An averaging period of 1 hour may not be appropriate if 

excursions of concentrations to greater than 1.5 times the 

average occur during the hour; in such cases, a shorter averaging 

period may be needed.) 

(2) the 4-day average concentration of u n -ionized ammonia 

(in mg/L NH3) does not exceed, more often than once every 3 years 

on the average, the average* numerical value given by 

0.80/FT/FPH/RATIO, where FT and FPH are as above and: 
RATIO = 16 : 7.7 - < pH - <9 

TCAP = 15 C; Salmonids or other sensitive 

coldwater species present 

coldwater species absent = 20 c; Salmonids and other sensitive 

(*Because these formulas are nonlinear in pH and temperature, the 

criterion should be the average of separate evaluations of the 

formulas reflective of the fluctuations of flow, pH, and 

temperature within the averaging period; it is not appropriate in 

general to simply apply the formula to average pH, temperature, 

and flow.) 

The extremes for temperature (0, 30) and pH (6.5, 9) given in 

the above formulas are absolute. It is not permissible with 

current data to conduct any extrapolations beyond these limits. 

In particular, there is reason to believe that appropriate 

criteria at pH > 9 will be lower than the plateau between pH 6 

and 9 given above. 

Criteria concentrations for the pH range 6.5 to 9.0 and the 

temperature range 0 C to 30 C are provided in the following 

tables. Total ammonia concentrations equivalent to each unionized ammonia concentration are also provided in these tables. 

There are limited data on the effect of temperature on chronic 

toxicity. EPA will be conducting additional research on the 

effects of temperature on ammonia toxicity in order to fill 

perceived data gaps. Because of this uncertainty, additional 

site -specif ic information should be developed before these 

0 
criteria are used in wasteload allocation modeling. For example, 

the chronic criteria tabulated for sites lacking salmonids are 

less certain at temperatures much below 20 C than those tabulated 

at temperatures near 20 C. Where the treatment levels needed to 

meet these criteria below 20 C may be substantial, use of sitespecific criteria is strongly suggested. Development of such 

criteria should be based upon site-specific toxicity tests. 

I) 

Data available for saltwater species are insufficient to 

derive a criterion for saltwater. 

The recommended exceedence frequency of 3 years is the 

Agency's best scientific judgment of the average amount of time 

it will take an unstressed system to recover from a pollution 

event in which exposure to ammonia exceeds the criterion. A 

stressed system, for example, one in which several outfalls occur 

in a limited area, would be expected to require more time for 

recovery. The resilience of ecosystems and their ability to 

recover differ greatly, however, and site-specific criteria may 

be established if adequate justification is provided. 

The use of criteria in designing waste treatment facilities 

requires the selection of an appropriate wasteload a1 location 

model. Dynamic models are preferred for the application of these 

criteria. Limited data or other factors may make their use 

impractical, in which case one should rely on a steady-state 

model. The Agency recommends the interim use of 1Q5 or lQlO for 

Criterion Maximum Concentration design flow and 745 or 7410 for 

the Criterion Continuous Concentration design flow in steadystate models for unstressed and stressed systems respectively. 
12) 4-day average concentrations for mmnla.. 

\ 

PU oc 5c I0 c 15 C 20 c 25 c WC 

L. SalmI6a or Other Smsltlv. Cold.ater SD~CI~S Present 

Un-1onlz.d Amonla lmg/lltr NH,) 

6.50 

6.75 

7 .w 

7.25 

7.50 

7.73 

8.00 

8.25 

8.50 

8.75 

9 .oo 

6.50 

6.75 

7 .oo 

7.25 

7.40 

7.75 

8.00 

8.25 

8.50 

8.75 

9 .oo 

0.0007 

0.0012 

0.0021 

0.0037 

0.0066 

0.0105 

0.0126 

0.0126 

O.Ol2t 

0.0126 

0.0126 

2.5 

2.5 

2.5 

2.5 

2.5 

2.3 

1.53 

0.87 

0.49 

0.28 

0.16 

0 .oow 

0.0017 

0.0029 

0.0052 

0.0093 

0.01 53 

0.0177 

0.0177 

0.0177 

0.0177 

0.0177 

2.4 

2.4 

2.4 

2.4 

2.4 

2.2 

I .44 

0.82 

0.47 

0.27 

0.16 

0.0013 

0.0023 

0.0042 

0.0074 

0.0132 

0.022 

0.025 

0.025 

0.025 

0.025 

0.025 

0.0019 

0.0033 

0 m59 

0.0105 

0.0186 

0.03 1 

0.035 

0.035 

0.035 

0.035 

0.035 

0.0019 

0.0033 

0.0059 

0.0105 

0.0186 

0.03 I 

0.035 

0.035 

0.035 

0.035 

0.035 

2.2 

2.2 

2.2 

2.2 

2.2 

2.1 

1.37 

0.70 

0.45 

0.26 

0.16 

2.2 

2.2 

23 

2.2 

2.2 

2 .o 

I .33 

0.76 

0.44 

0.27 

0.16 

1 A9 

I ,49 

1.49 

I .50 

I .% 

I .40 

0.93 

0.54 

0.32 

0.19 

0.13 

0.0019 

0.0033 

0.00% 

0.0105 

0.0186 

0.031 

0 .a35 

0.035 

0.035 

0.035 

0.035 

I .M 

1.04 

1 .04 

I .M 

1.05 

0.99 

0.66 

0.39 

0.23 

.. 0.15 

0.10 

0.0019 

0.0033 

0.0059 

0.0105 

0.0186 

0.031 

-0.035 

0.035 

0.035 

0.035 

0.035 

0.73 

0.73 

Q.74 

0.74 

0.74 

0.71 

0.47 

0.28 

0.17 

0.11 

0 .oc 

8 ;oo 0;0126 

8.25 0.0126 

8.50 0.0126 

8.75 0.0120 

9.00 0.0126 

8.50 o;r9 

8.75 0.28 

9 .oo 0.16 

0.0009 

0.0017 

0.0029 

0.0052 

0.0093 

0.0153 

0.0177 

0.0177 

0.0177 

0.0177 

0.0177 

0.0013 

0.0023 

0.0042 

0.0074 

0.0132 

0.022 

0.025 

0.025 

0.025 

0.025 

0.025 

0.0019 

0.0033 

0.0059 

0.0105 

0.0186 

0.03 1 

0.035 

0.035 

0.035 

0.035 

0.035 

0.0026 

0.0047 

0 .Og83 

0.0148 

0.026 

0.043 

0.0% 

0.050 

0.0% 

0.050 

0.050 

Total Amonla (mg/lltor NH,) 

2.1 

2. I 

2.1 

2. I 

2.1 

I .98 

1.51 

0.76 

c.45 

0.27 

0.17 

0 .0026 

0.0047 

0.0083 

0.0148 

0.026 

0.043 

0.0% 

0.050 

0.0% 

0.0% 

o.om 

I .a 

I A7 

I A7 

I .48 

1 A9 

1.39 

0.93 

0.54 

0.33 

0.21 

0.14 

0.W26 

0.0047 

0.0083 

0.0148 

0.026 

0.043 

0.050 

0.050 

0.050 

0.050 

0.050 

I .03 

1-04 

1.04 

1.05 

I .06 

I .oo 

0.67 

0.40 

0 25 

0.16 

0.11 

To convert meso valws to np/ll?er N, multlply by 0.822. 

t Sltc~p~~ltIc crltrla dorolopment Is rtronplv ruggost.d at TmDOratyr" abwm 20 C 

buause of the Iln1t.d data avallable to generato tho crltula rrm.nd.tlon, and 

at tmperatur*s blow 20 C bWuse of Tb lim1t.d data and boauY mall hmgor In 

the crltorla may have rlgnlflcant Impact on the Inol Of trmtl(lt r.qu1r.l In 

Irotlng tho r.6umNnd.d crltorla. 
oc 5c 10 c I5 c MC 25 C KJC vH 

A. Salmcnldr or Omer Sensltlve CoId.atu 5pul.r Present 

6.50 35 

6.75 32 i ;oo 28 

7 -25 23 

7.50 17.4 

7.75 12.2 

8 .oo 8 .o 

8.25 4.5 

8.30 . 2.6 

8.75 I .47 

9 .oo 0.86 

Vn-lonlz.d Amanla 

0.0129 0.0182 

0.02 1 0.030 

0.033 0 .w 

0.048 0.W8 

0 .ow 0.091 

o.oa0 0,113 

0.092 0.133 

0.092 0.130 

0.092 0.130 

0.092 0.130 

0.092 0.1x 

(mq/llter 

0.026 

0.042 

0 .W6 

0.095 

0.128 

0.159 

0.1W 

0.184 

0.lW 

0.184 

0.184 

NH3) 

0.036 

0.059 

0.093 

0.135 

0.181 

0.22 

0 26 

0.26 

0 26 

0.26 

0 36 

33 

33 

26 

22 

I63 

11.4 

7.5 

4.2 

I .40 

0.83 

2.4 

31 

28 

25 

20 

15.5 

10.9 

7 .I 

4.1 

2.3 

I .37 

0 .83 

30 

27 

24 

19.7 

14.9 

10.5 

6.9 

2.3 

I .)8 

0 36 

4.0 

29 

27 

23 

19.2 

14.6 

10.3 

6 .8 

3.9 

2.3 

1.42 

0.91 

0.0s 

0.059 

0.093 

0.135 

0.101 

0.22 

0.26 

0.26 

0.26 

0.26 

0.26 

20 

18.6 

16.4 

13.4 

10.2 , 

7 .2 

4.8 

2.8 

1.71 

1.07 

0.72 

0.036 

0.059 

0.093 

0.135 

0.181 

0.22 

0 36 

0.26 

0.26 

0.26 

0 26 

14.3' 

13.2 

11.6 

9.5 

7.3 

5.2 

3.5 

2.1 

1.28 

0.83 

0 .58 

6.50 

6.75 

7 .w 

7 ;25 

7.w 

7.75 

8 .oo 

8.25 

8 :59 

8.75 

9 .oo 

6.50 

6.75 

1.00 

7.25 

7.50 

7 .7) 

8 .oo 

0.25 

8.50 

8.75 

9 .w 

0.009 I 

0.0149 

0.023 

0.034 

0.045 

0.056 

0.065 

0.065 

0.065 

0.065 

0.065 

35 

32 

28 

2) 

17 A 

12.2 

8 .o 

4.5 

2.6 

I .41 

0.86 

0.0129 

0.02 I 

0.033 

0.046 

0 ;OM 

0.080 

0 .ox 

0.092 

0.092 

0.092 

0 .on 

0.0182 

0.030 

0.046 

0.068 

0.091 

0.113 

0.130 

0.130 

0.130 

0.150 

0.130 

3s 

30 

26 

22 

16.3 

4.2 

2.4 

I .40 

0.83 

0.026 

0.042 

0.066 

0.095 

0.128 

0.159 

0.184 

0.184 

0.184 

0.184 

0.184 

0.036 

0.059 

0.093 ~.~ ~ 

0.135 

0.181 

0.22 

0.26 

0.26 

0;26 

0.26 

0.26 

31 

28 

25 

20 

15.5 

10.9 

7 .I 

4;l 

2.3 

1.31 

0.83 

30 

27 

24 

19.7 

14.9 

10.5 

6.9 

4.0 

2 i3 

1.33 

0.86 

29 

27 

23 

19.2 

14.6 

10.3 

6.8 

3.9 

2.3 

1.42 

0.91 

0.051 

0.084 

0.131 

0.190 

0 26 

0.32 

0.37 

0.31 

0.37 

0.37 

0 37 

29 

26 

23 

19.0 

14.5 

10.2 

6.8 

4 .O 

2 ;4 

1.52 

1.01 

0.051 

0.131 

0.190 

0.26 

0.32 

0.37 

0.37 

0.37 

0.37 

0.37 

0.084 

20 

18.6 

16.4 

13.5 

10.3 

7.3 

4.9 

2;9 

1 .el 

1.18 

0.82 
The Agency acknowledges that the Criterion Continuous 

Concentration stream flow averaging period used for steady -state 

wasteload allocation modeling may be as long as 30 days in 

situations involving POTWs designed to remove ammonia where 

limited variability of effluent pollutant concentration and 

resultant concentrations in receiving waters can be demonstrated. 

In cases where low variability can be demonstrated, longer 

averaging periods for the ammonia Criterion Continuous 

Concentration (e.g., 3 0 - day averaging periods) would be 

acceptable because the magnitude and duration of exceedences 

above the Criterion Continuous Concentration would be 

sufficiently limited. These matters are discussed in more detail 

in the Technical Support Document for Water Quality-Based Toxics 

Control (U.S. EPA, 1985a). 

(50 F.R. 30784, July 29, 1985) 

SEE APPENDIX A FOR METHODOLOGY 
BERYLLIUM 

CRITERIA: 

8 Aquatic Life 

The available data for beryllium indicate that acute and 

chronic toxicity to freshwater aquatic life occur at 

concentrations as low as 130 and 5.3 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. Hardness has a substantial effect 

on acute toxicity. 

The limited saltwater data base available for beryllium does 

not permit any statement concerning acute or chronic toxicity. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to beryllium through ingestion 

of contaminated water and contaminated aquatic organisms, the 

ambient water concentration should be zero, based on the non 

threshold assumption for this chemical. However, zero level may 

not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

lifetime are estimated at and The 

corresponding recommended criteria are 68 ng/L, 6.8 ng/L, and 

0.68 ng/L, respectively. If these estimates are made for 
consumption of aquatic organisms only, excluding consumption of 

water, the levels are 1170 ng/L, 117.0 nq/L, and 11.71 ng/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 

a 
BORON 

CRITERION: 

750 ug/L for long-term irrigation on sensitive crops. 

INTRODUCTION: 

Boron is not found in its elemental form in nature: it is 

usually found as a sodium or calcium borate salt. Boron salts 

are used in fire retardants, the production of glass, leather 

tanning and finishing industries, cosmetics, photographic 

materials, metallurgy and for high energy rocket fuels. 

Elemental boron also can be used in nuclear reactors for neutron 

absorption. Borates are used as "burnable" poisons. 

RATIONALE: 

Boron is an essential element for growth of plants butthere 

is no evidence that it is required by animals. The maximum 

concentration found in 1,546 samples of river and lake waters 

from various parts of the United States was 5.0 mg/L: the mean 

value was 0.1 mg/L (Kopp and Kroner, 1967). Ground waters could 

contain substantially higher concentrations at certain places. 

The concentration in seawater is reported as 4.5 mg/L in the forn 

of borate (NAS, 1974). Naturally occurring concentrations of 

boron should have no effects on aquatic life. 

The minimum lethal dose for minnows exposed to boric acid at 

20 OC for 6 hours was reported to be 18,000 to 19,000 mg/L in 

distilled water and 19,000 to 19,500 mg/L in hard water (Le Clerc 

and Devlaminck, 1955: Le Clerc, 1960). 

In the dairy cow, 16 to 20 g/day of boric acid for 40 days 
produced no ill effects (McKee and Wolf, 1963). 

Sensitive crops have shown toxic effects at 1000 ug/L or 

less of boron (Richards, 1954). Bradford (1966), in a review of 

boron deficiencies and toxicities, stated that when the boron 

concentration in irrigation waters was greater than 0.75 ug/L, 

some sensitive plants such as citrus began to show injury. 

Biggar and Fireman (1960) showed that with neutral and alkaline 

soils of high absorption capacities, water containing 2 ug/L 

boron might be used for some time without injury to sensitive 

plants. The criterion of 750 ug/L is thought to protect 

sensitive crops during long-term irrigation. 

(QUALITY CRITERIA €OR WATER, JULY 1976) PB-263943 

SEE APPENDIX C €OR METHODOLOGY 
CHLORINATED BENZENES 

CRITERIA e Aquatic Life 

The available data for chlorinated benzenes indicate that 

acute toxicity to fresh water aquatic life at concentrations as 

low as 250 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data are 

available concerning the chronic toxicity of the more toxic of 

the chlorinated benzenes to sensitive freshwater aquatic life but 

toxicity occurs at concentrations as low as 50 ug/L for a fish 

species exposed for 7.5 days. 

The available data for chlorinated benzenes indicate that 

acute and chronic toxicity to saltwater aquatic life occur at 

concentrations as low as 160 and 129 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

Human Health 

For comparison purposes, two approaches were used to derive 

criterion levels for monochlorobenzene. Based on available 

toxicity data, for the protection of public health, the derived 

level is 488 ug/L. Using available organoleptic data, for 

controlling undesirable taste and odor quality of ambient water, 

the estimated level is 20 ug/L. It should be recognized that 

organoleptic data as a basis for establishing a water quality 

criteria have limitations and have no demonstrated relationship 

to potential adverse human health effects. 
Trichlorobenzenes 

Due to the insufficiency in the available information for the 

trichlorobenzenes, a criterion cannot be derived at this time 

using the present guidelines. 

1,2,4,5-Tetrachlorobenzene 

For the protection of human health from the toxic properties 

of 1,2,4,5-tetrachlorobenzene ingested through water and 

contaminated aquatic organisms, the ambient water criterion is 

determined to be 38 ug/L. 

For the protection of human health from the toxic properties 

of 1,2,4,5-tetrachlorobenzene ingested through contaminated 

aquatic organisms alone, the ambient water criterion is 

determined to be 48 ug/L. 

Pentachlorobenzene 

For the protection of human health from the toxic properties 

of pentachlorobenzene ingested through water and contaminated 

aquatic organisms, the ambient water criterion is determined to 

be 74 ug/L. 

For the protection of human health from the toxic properties 

of pentachlorobenzene ingested through contaminated aquatic 

organisms alone, the ambient water criterion is determined to be 

85 ug/L. 

Hexachlorobenzene 

For the maximum protection of human health from the potential 

carcinogenic effects due to exposure of hexachlorobenzene through 

ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentration should be zero based 

on the non-threshold assumption for this chemical. However, zero 
level may not be attainable at the present time. Therefor, the 

levels which may result in incramental increase of cancer risk 

0 over the lifetime are estimated at and The 

corresponding recommended criteria are 7.2 ng/L, 0.72 ng/L, and 

0.072 ng/L, respectively. If the above estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 7.4 ng/L, 0.74 ng/L and 0.074. ng/L 

respectively. 

(45 F.R. 79316, November 28, 1980) 

SEE APPERDIX B FOR METHODOLOGY 
DICHMROPROPANES/DICHMROPROPENES 

CRITERIA: 

Aquatic Life 

The available data for dichloropropanes indicate that acute 

and chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 23,000 and 5,700 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for dichloropropene indicate that acute 

and chronic toxicity to freshwater aquatic life occurs at 

concentrations as low as 6,060 and 244 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for dichloropropane indicate that acute 

and chronic toxicity to saltwater aquatic life occur at 

concentrations as low as 10,300 and 3,040 ug/L, respectively, and 

would occur at lower concentrations among species that are more 

sensitive than those tested. 

The available data for dichloropropene indicate that acute 

toxicity to saltwater aquatic 1 ife occurs at concentrations as 

low as 790 ug/L and would occur at lower concentrations among 

species that are more sensitive than those tested. No data 

are available concerning the chronic toxicity of 

dichloropropene to sensitive saltwater aquatic life. 
Human Health 

Using the present guidelines, a satisfactory criterion cannot 

be derived at this time because of insufficient avaiiable data 

for dichloropropanes. 

For the protection of human health from the toxic properties 

of dichloropropenes ingested through water and contaminated 

aquatic organisms, the ambient water criterion is determined to 

be 87 ug/L. 

For the protection of human health from the toxic properties 

of dichloropropenes ingested through contaminated aquatic 

organisms alone, the ambient water criterion is determined to be 

14.1 mg/L. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B €OR METHODOLOGY 
CRITERIA: 

- 

For endrin the crit 

*ENDRIN 

Aquatic Life 

rion to protect fr shwater aquatic li e 

as derived using the Guidelines is 0.0023 ug/L as a 24-hour 

average, and the concentration should not exceed 0.18 ug/L at any 

time. 

For endrin the criterion to protect saltwater aquatic life as 

derived using the Guidelines is 0.0023 ug/L as a 24-hour average, 

and the concentration should not exceed 0.037 ug/L at any time. 

Human Health 

The ambient water quality criterion for endrin is recommended 

to be identical to the existing water standard which is 1.0 ug/L. 

Analysis of the toxic effects data resulted in a calculated level 

which is protective of human health against the ingestion of 

contaminated water and contaminated aquatic organisms. The 

calculated value is comparable to the present standard. For 

this reason a selective criterion based on exposure solely from 

consumption of 6.5 g of aquatic organisms was not derived. 

*Indicates suspended, canceled or restricted by U.S. EPA Office 

of Pesticides and Toxic Substances 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
HEPTACHMR 

CRITERIA: 

Aquatic Life 0 For heptachlor the criterion to protect freshwater aquatic 

life as derived using the Guidelines is 0.0038 ug/L as a 24-hour 

average, and the concentration should not exceed 0.52 ug/L at any 

time. 

For heptachlor the criterion to protect saltwater aquatic 

life as derived using the Guidelines is 0.0036 ug/L as a 24-hour 

average, and the concentration should not exceed 0.053 ug/L at 

any time. 

Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to heptachlor through ingestion 

of contaminated water and contaminated aquatic organisms, the 

ambient water concentration should be zero, based on the non 

threshold assumption for this chemical. However, zero level 

may not be attainable at the present time. Therefore, the levels 

which may result in incremental increase of cancer risk over the 

lifetime are estimated at and The 

corresponding recommended criteria are 2.78 ng/L, 0.28 ng/L, and 

0.026 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 2.85 ng/L, 0.29 ng/L, and 0.029 ng/L, 

respectively. 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
CRITERIA : 

HE?ACHLOROCYCU3HEXANE 

Aquatic Life 

€or lindane the criterion to protect freshwater aquatic life 

as derived using the Guidelines is 0.080 ug/L as a 24-hour 

average and the concentration should not exceed 2.0 ug/L at any 

time. 

For saltwater aquatic life the concentration of lindane 

should not exceed 0.16 ug/L at any time. No data are available 

concerning the chronic toxicity of lindane to sensitive saltwater 

aquatic life. 

BHC 

The available data for a mixture of isomers of BHC indicate 

that acute toxicity to freshwater aquatic life occurs at 

concentrations as low as 100 ug/L and would occur at lower 

concentrations anong species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity of 

a mixture of isomers of BHC to sensitive freshwater aquatic life. 

The available data for a mixture of isomers of BHC indicate 

that acute toxicity to saltwater aquatic life occurs at 

concentrations as low as 0.34 ug/L and would occur at lower 

concentrations among species that are more sensitive than those 

tested. No data are available concerning the chronic toxicity of 

a mixture of isomers of BHC to sensitive saltwater aquatic life. 
Human Health 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to alphahexachlorocyclohexane through ingestion of contaminated water and 

contaminated aquatic organisms, the ambient water concentrations 

should be zero, based on the nonthreshold assumption for this 

chemical. However, zero level may not be attainable at the 

present time. Therefore, the levels which may result in 

incremental increase of cancer risk over the lifetime are 

estimated at 10-6 and The corresponding 

recommended criteria are 92 ng/L, 9.2 ng/L, and .92 ng/L, 

respectively. If these estimates are made for consumption of 

aquatic organisms only, excluding consumption of water, the 

levels are 310 ng/L, 31.0 ng/L, and 3.10 ng/L, respectively. 

For the maximum protection ofrhuman health from the potential 

carcinogenic effects of exposure to beta-hexachlorocyclohexane 

through ingestion of contaminated water and contaminated aquatic 

organisms, the ambient water concentrations should be zero, based 

on the nonthreshold assumption for this chemical. However, zero 

level may not be attainable at the present time. Therefore, thc 

levels which may result in incremental increase of cancer risk 

over the lifetime are estimated at lo+, and 

The corresponding recommended criteria are 163 ng/L, 16.3 ng/L, 

and 1.63 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 547 ng/L, 54.7 ng/L, and 5.47 ng/L, 

respectively. 
For the maximum protection of human health from the potential 

carcinogenic effects due to exposure of gamahexachlorocyclohexane through ingestion of contaminated water and 

contaminated aquatic organisms, the ambient water concentrations 

should be zero, based on the nonthreshold assumption for this 

chenical. However, zero level may not be attainable at the 

present time. Therefore, the levels which may result in 

incremental increase of cancer risk over the lifetime are 

estimated at lom5, and The corresponding 

recommended criteria are 186 ng/L, 18.6 ng/L, and 1.86 ng/L, 

respectively. If these estimates are made for consumption of 

aquatic organisms only, excluding consumption of water, the 

levels are 625 ng/L, 62.5 ng/L, and 6.25 ng/L, respectively. 

For the maximum protection of human health from the potential 

carcinogenic effects of exposure to technicalhexachlorocyclohexane through ingestion of contaminated water 

and contaminated aquatic organisms, the ambient water 

concentrations should be zero, based on the nonthreshold 

assumption for this chemical. However, zero level may not be 

attainable at the present time. Therefore, the levels which may 

result in incremental increase of cancer risk over the 

lifetime are estimated at and The 

corresponding recommended criteria are 123 ng/L, 12.3 ng/L, and 

1.23 ng/L, respectively. If these estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 414 ng/L, 41.4 ng/L, and 4.14 ng/L, 

respectively. 

0 
Using the present guidelines, satisfactory criteria cannot be 

derived at this tine for delta and epsilon hexachlorocyclohexane 

because of insufficient available data. a 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
MIREX 

CRITERION: 

0.001 ug/L for freshwater and marine aquatic life. 0 

RATIONALE: 

Mirex is used to control the imported fire ant Solenssis 

saevissima richteri in the southeastern United States. Its use 

is essentially limited to the control of this insect and it is 

always presented in bait. In the most common formulation, 

technical grade mirex is dissolved in soybean oil and sprayed on 

corncob grits. The bait produced in this manner consists of 0.3 

__ 

percent mirex, 14.7 percent soybean oil and 85 percent corncob 

grits. bait often is applied at a rate of 1.4 kg/ha, 

equivalent to 4.2 grams of toxicant per hectare. 

The mirex 

Relatively few studies have been made of the effects of mirex 

on freshwater invertebrates of 'these, only Ludke et al. (1971) 

report chemical analyses of mirex in the water. Their study 

reported effects on two crayfish species exposed to mirex by 

three techniques. First, field - collected crayfish were exposed 

to several sublethal concentrations of technical grade mirex 

solutions for various periods of time: second, crayfish were 

exposed to mirex leached from bait (0.3 percent active 

ingredient) ; and third, the crayfish were fed mirex bait. 

Procambarus - blandingi juveniles were exposed to 1 or 5 ug/L 

for 6 to 144 hours, transferred to clean water and observed for 

10 days. After 5 days in clean water, 95 percent of the animals 

exposed to 1 ug/L for 144 hours were dead. Exposure to 5 ug/L 

for 6, 24, and 58 hours resulted in 26, 50, and 98 percent 

mortality 10 days after transfer to clean water. Crayfish, 
Procambarus w, were exposed to 0.1 and 0.5 ug/L for 48 hours. 

Four days after transfer to clean water, 65 percent of the 

animals exposed to 0.1 ug/L were dead. At the 0.5 ug/L 

concentration, 71 percent of the animals were dead after 4 days 

in clean water. Tissue residue accumulations (wet weight basis) 

ranged from 940 - to 27,210-fold above water concentrations. In 

leached bait experiments, 10 bait particles were placed in 2 

liters of water but isolated from 20 juvenile crayfish. Thirty 

percent of the crayfish were dead in 4 days and 95 percent were 

dead in 7 days. Water analysis indicated mirex concentrations of 

0.86 ug/L. In feeding experiments, 108 crayfish each were fed one 

bait particle. Mortality was noticed on the first day after 

feeding, and by the sixth day 77 percent were dead. In another 

- 

* 

experiment, all crayfish were dead 4 days after having been fed 2 

bait particles each. From this report it is obvious that mirex is 

extremely toxic to these species of crayfish. Mortality and 

accumulation increase with time of exposure to the insecticide. 

Concentrations as low as 0.1 ug/L or the ingestion of one 

particle resulted in death. 

Research to determine effects of mirex on fish has been 

concentrated on species which have economic and sport fishery 

importance. Hyde et al. (1974) applied mirex bait (0.3 percent 

mirex) at the standard rate (1.4 kg/ha) in four ponds containing 

channel catfish, - Ictalurus punctatus. Three applications were 

made over an 8-month period with the first application 8 days 

after fingerling (average weight 18.4 g) catfish were placed in 

the ponds. Fish were collected at each subsequent application 

- 

* 
(approximately 4-month intervals). Two and one half months after 

the final application, the ponds were drained, all fish were 

measured and weighed, and the percent survival was calculated. 

Mirex residues in the fish at termination of the experiment 

ranged from 0.015 ug/g (ppm) in the fillet to 0.255 ug/g in the 

fat. 

0 

In another study, Van Valin et al. (1968) exposed bluegills, 

- LeEmis -_ macrochirus, -- and the goldfish, Carassius auratus, to 

mirex by feeding a mirex-treated diet (1, 3, and 5 mg mirex per 

kg body weight) or by treating holding ponds with mirex bait 

(1.3, 100, and 1000 ug/L computed water concentration). They 

reported no mortality or tissue pathology for the bluegills: 

however, after 56 days of exposure, gill breakdown in goldfish 

was found in the 100 and 1000 ug/L contact exposure ponds, and 

kidney breakdown was occurring in the 1000 ug/L ponds. Mortality 

in the feeding experiments was not related to the level of 

exposure, although growth of the bluegills fed 5 ug/L mirex was 

reduced. 

In laboratory and field test systems, reported concentrations 

of mirex usually are between 0.5 and 1.0 ug/L (Van Valin et al. 

1968: Ludke et al. 1971). Although mirex seldom is found above 1 

ug/L in the aquatic environment, several field studies have shown 

that the insecticide is accumulated through the food chain. 

Borthwick et al. (1973) reported the accumulation of mirex in 

South Carolina estuaries. Their data revealed that mirex was 

transported from treated land and marsh to the estuary animals 

and that accumulation, especially in predators, occurred. In the 

test area, water samples consistently were less than 0.01 ug/L. 
Residues in fish varied from non - detectable to 0.8 ug/g with 15 

percent of the samples containing residues. The amount of mirex 

and the percent of samples containing mirex increased at higher 

trophic levels. Fifty -four percent of the raccoons sampled 

contained mirex residues up to 4.4 ug/g and 78 percent of the 

birds contained residues up to 17 ug/g. Nagvi and de la Cruz 

(1973) reported average residues for molluscs (0.15 ug/g), fish 

(0.26 ug/g), insects (0.29 ug/g), crustaceans (0.44 ug/g) and 

annelids (0.63 ug/g). They also reported that mirex was found 

in areas not treated with mirex which suggests movement of the 

pesticide in the environment. Wolfe and Norment (1973) sampled 

an area for one year following an aerial application of mirex 

bait (2.1 g mirex/ha). Crayfish residues ranged from 0.04 to 

0.16 ug/g. Fish residues were about 2 to 20 times greater than 

the controls and averaged from 0.01 to 0.76 ug/g. Kaiser (1974), 

reported the presence of mirex in fish from the Bay of Quinte, 

Lake Ontario, Canada. Concentrations range from 0.02 ug/g in 

the gonads of the northern long nose gar, Lepistosteus osseus, to 

0.05 ug/g in the post- anal fin of the northern pike, Esox lucius. 

Mirex has never been registered for use in Canada. 

Mirex does not appear to be greatly toxic to birds, with 

LC5O's for the young of four species ranging from 547 to greater 

than 1667 ug/g (Heath et al. 1972). Long-term dietary dosages 

caused no adverseeffect at3 ug/gwithmallards and 13 ug/gwith 

pheasants (Heath and Spann, 1973). However, it has been reported 

(Stickel et al. 1973) that the persistence of mirex in bird 

tissue exceeds that of all organochlorine compounds tested except 
for DDE. Delayed mortality occurred among birds subjected to 

doses above expected environmental concentration. 

A summary examination of the data available a*this time 

shows a mosaic of effects. Crayfish and channel catfish survival 

.is affected by mirex in the water or by ingestion of the bait 

particles. - Bioaccumulation is well established for a wide 

variety of organisms but the effect of this bioaccumulation on 

the aquatic ecosystem is unknown, There is evidence that mirex 

is very persistent in bird tissue. Considering the extreme 

toxicity and potential for bioaccumulation, every effort should 

be made to keep mirex bait particles out of water containing 

aquatic organisms and water concentrations should not exceed 

0.001 uq/L mirex. This value is based upon an application factor 

of.0.01 applied to the lowest levels at which effects on crayfish 

have been observed. 

Data upon which to base a marine criterion involve several 

estuarine and marine crustaceans. A concentration of 0.1 ug/L 

technical grade mirfx in flowing seawater was lethal to juvenile 

pink shrinp, Penaeus durorarum, in a 3-week exposure (Lowe et al. 

1971). In static tests with larval stages (megalopal) of the mud 

crab, Rhithropanopeus harrisii, reduced survival was observed in 

B.l-ug/L mirex (Bookhout et al. 1972). In three of four 28-day 

seasonal flow-through experiments, Tagatz et al. (1975) found 

Educed survival of Call inectes sapidus, Penaeus durorarum, and 

grass shrimp, Palaemonetes pugio, at levels of 0.12 ug/L in 

summer, 0.06 ug/L in fall and 0.09 ug/L in winter. 

Since two reports, Lowe et al. (1971) and Bookhout et al. 

(1972), stated that effects of mirex on estuarine and marine 
crustaceans were observed only after considerable time had 

elapsed, it seems reasonable that length of exposure is an 

important consideration for this chemical. This may not be the 

case in fresh water since the crayfish were affected within 48 

hours. Therefore, a 3- to 4-week exposure night be considered 

tlacutetB and by applying an application factor of 0.01 to a 

reasonable average of toxic-effect levels as summarized above, a 

recommended marine criterion of 0.001 ug/L results. 

(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943 

SEE APPENDIX C FOR METHODOMGY 
NICKEL 

CRITERIA: 

Aquatic Life 

For total recoverable nickel the criterion (in ug/L) to 

0 

protect freshwater aquatic life as derived using the Guidelines 

is the numerical value given by e(0.76[ln(hardness)]+l.06) as a 

24- hour average, and the concentration (in ug/L) should not 

exceed the numerical value given by .(0.76[ In (hardness) ]+4.02) 

at any time. For example, at hardnesses of 50, 100, and 200 mg/L 

as CaC03 the criteria are 56, 96, and 160 ug/L, respectively, as 

24-hour averages, and the concentrations should not exceed 1,100, 

1,800, and 3,100 ug/L, respectively, at any time. 

For total recoverable nickel the criterion to protect 

saltwater aquatic life as derived using the Guidelines is 7.1 

ug/L as a 24-hour average, and'the concentration should not 

0 exceed 140 ug/L at any time. 

Human Health 

For the protection of human health from the toxic properties 

of nickel ingested through water and contaminated aquatic 

organisms, the ambient water criterion is determined to be 13.4 

W/L. 

For the protection of human health from the toxic properties 

of nickel ingested through contaminated aquatic organisms 

alone, the ambient water criterion is determined to be 100. 

Ug/L* 

(45 F.R. 79318, November 28, 1980) 

SEE APPENDIX B FOR METHODOLOGY 
2,3,7,8-TETRACHLORODIBENZO-P-DIOXI€l 

CRITERIA: 

Aquatic Life a Not enough data are available concerning the effects of 

2,3,7,8-TCDD on aquatic life and its uses to allow derivation of 

national criteria. The available information indicates that 

acute values for some freshwater animal species are >1.0 ug/L; 

some chronic values are <0.01 ug/L; and the chronic value 

for rainbow trout is <0.001 ug/L. Because exposures of 

some species of fishes to 0.01 ug/L for <6 days resulted in 

substantial mortality several weeks later, derivation of 

aquatic life criteria for 2,3,7,8-TCDD may require special 

consideration. Predicted bioconcentration factors (BCFs) for 

2,3,7,8-TCDD range from 3,000 to 900,000, but the available 

measured BCFs range from 390 to 13,000. If the BCF is 5,000, 

concentrations >0.00001 ug/L should result in concentrations 

in edible freshwater and saltwater fish and shellfish that 

exceed levels identified in a US. FDA health advisory. If the 

BCF is >5,000 or if uptake in a field situation is greater than 

that in laboratory tests, the value of 0.00001 ug/L will be too 

high. 

0 

Human Health 

€or the mxirnux protection of human health from the potential 

carcinogenic effects of 2,3,7,8-TCDD exposure through ingestion 

of contaminated water and contaminated aquatic organisms, the 

ambient water concentration should be zero. This criterion is 
based on the nonthreshold assumption for 2,3,7,8-TCDD. However, 

zeromay not be an attainable level at this time. Therefore, the 

levels that may result in an increase of cancer risk over the 

lifetime are estimated at and The 

corresponding recommended criteria are 1.3~10-~, 1.3~10-~ and 

1.3x10-’ ug/L, respectively. If the above estimates are made for 

consumption of aquatic organisms only, excluding consumption of 

water, the levels are 1.4x30-~, 1.4x10-* and 1.4x10-’ ug/L, 

respectively. If these estimates are made for comsumption of 

water only, the levels are 2.2~10-~, 2.2xlO-’ and 2.2~10-~ ug/L, 

respectively. 

(49 F.R. 5831, February 15, 1984) 

SEE APPENDIX B €OR METHODOLOGY 
ALKALINITY 

BARIUM 

REFERENCES CITED 

Browning, E 1961. Toxicity of industrial metals. Butterworth, London. 

btr. M., et d. 1970. Effeeta of pollution on fmh Me, heavy mea Annd Literature 

review. Jw. Water Poll. Cont Fed. 42W. 

L.ngr. NA 1961. Handbook of cbemistry. lorh ed Yd;fiw-Hill, Book Co.. New Ywk. 

little, A.D. 1971. Iwrg.oic cbemical pollution of fnah water. Water qdty &la lmk 

Val. 2. US. Envimnmental Proteetion Agency, la010 DPV. pp. 2e26. 

McKee, J.E. ud W.W. Wolf. 1963. Water quality criteria Calif& Skte Watm 

Muma Contrd Board, Pub. No. 3-A 

National Aademy of Seieooes. National Academy of Engioeering. 191+ Water qwlity 

criteria, 1972 US Government Printing Off=, W.shin@n, D.C. 

Patty, FAl962lndustrialbygieneudtoximlogy,VoL1I.JohnA Wiley,NerYatpp 

%loo2 Cited in US. Department of Health, Eduatian ud Welfare. 1M. 

Public Health SeMce. 1WB. Drinking water quality of rkasd intast.te eUria 

water supplies. U.S. Depar(ment of Health Eduatian ud Welfy W..hin%on, D.C. 

Sollmann. T.H. 1957. A manud of pharmamlogy and its applition to ud 

toljcolcgy. 8th ed. W.B. Saunden Co. Philadelphia 

Stokinger, H.E.,radRLWmdwaardl968.Toljcdogiemetbodrfaat.blirhin~ 

water standards. Jour. Ame?. Water Worlrs h. 50.615. 

U.S Deprtment of Health, Education and Welfue. 1968. Preliminvy air pdlutioa 

rwey of buiurn and its compouinja, a literature review. NaW Air PoUdn CODtml 

Administntion Publiation No. AIT D 69-23, Raleigh, N.C. 

BORON 

REFERENCES CITED 

Bw. J.W.. and M. Fireman 1960. Boron *tion and mlean~ by mik Soil sd. SOe 

Amer.Ra2-l: 115. 

Bdwd. G.R 1966. hn [toxicity, indicator plants], in di.gwstic criteria for plants 

and .oik H.D. Chapman. ed Univwsity of California, Diviaion of Agridtural Science, 

Berkeley, p. 53. 

Le Clem E 1960. The self-purifkation of strams ud the relatia~hip between chemical 

md biologid test& Ra Seeond Symposium on Treatment of Want? Watem. 

Pergamon h. London, p. 231. 

Le Clem. E, and F. DevhindL 1456. Fiah toxicity tats and water qdty. Bull. de Beige 

Condument E.u. 28: 11. 

Kopp J.F.. ud RC. Kmner. 1961. Trace met& in waten of the UnitEd Slate.. Fed4 

Water Pollution Conbol Administration, US. Department of Interior, Cincinnati. Ohio. 

YcKee, J.E, ud A.W. Wolf. 1963. Water quality criteria State Water Qrulity Conmi 

Bcwd. haamento, Calif. Pub. 3-A. 

National Ardemy of Seieocpa, National Academy of Engineering. 1974 Water pdity 

mi- 1972 US. Government Printing Offii. W.shingon, D. C. 

RiW. LA. ed. 1954. Diagnosis and impmvement of saline md Iwi mi. AgrieultHmdbmk No. €U. US. Government F’rinting Offioe, Wpshington, D. C. 
CHMROPHENOXY HERBICIDES 

REFERENCES CITED 

Knua. u dted by Mitcheaa. J.W., RE Hogmn, and C.R. Gaetjem. 1946. Tol- of 

fum .ni& to feed mnUining 2,4diehlomphenoxyacetie db Jour. Anid. sd. 

5:226. 

Lthrmn. AJ. 1986. Sum- of pticide toxidty. Assodrtbn of Food md Dnrg 

Oificialn of the US.. Toph KAM. pp. 1-40, 

Palmer. J.S.. and RD. Radeleff. 1964. The toximlcgk elf& of certain fungiddea md 

herbicida on sheep and cattle. Ann. N.Y. Ad. Sei. 111: 729. 

Wury, J.H. 1963 Toxicity of 24dichloruphenoxyacetic acid for man and dog. Arch. 

Envir Health. 7: 282 

COLOR 

REFERENCES CITED 

hceriom Water Works Assodrtion. 1971. Water qudity and treatment Brd ed MdmwBirge. EX. md C. Juday. 1990. A amnd report on .oh radiation and inland Up.. 

BU A.P., and RF. chrisrmaa l96sachvrrtMstw . of colored surface waters Jau. 

Hill Book CO., New York 

Tnns. Wk. Ad. Science, AM. Let. a: 285. ' 

hmer. Water works kdn. 55: 7s. 

El&. AP.. and RF. chrisrmaa 1953b. Cbemi~I rby.ctaufKI ' of fdvk dda Jau. ' 

~~ 

Akner.w.7eTworksAasa,55:w. 

Natio~l Ademy of Mencm National Academy of Engirraing. 1974 Wsta qdty 

N~tio~l Technical Adviawy Committee to tbe Secreq of tbe IntpiOr. 1988. Wata 

Public iiealth service. 1952 Drinkira water st.nduda, 1962 PAS F'uM No. SS6. US. 

criterir 1972 US. Government Rinting Office. Wnhingmn, D.C. 

quality criteria US. Government Printing Office, Wuhingron, D.C. 

Government Printing oifii, Washir;gOn. D.C. 

Sawyer. C.N. 1960. Chemistry for ..nitary engineera. Mdmw-Hill Bodr Co., New YOrL. 

Shn~w. J. lSi Effect of yellow ode dds on Lon md other meub in water. Jour. 

A'-: Water works hti 56: 1062' 

Stllldvd methods for the examination of water and wastewater. l8tb ed. 1971. Edited by 

Michael C. Rand. et d. Moan Public Health Aran. American Wster Works kdn. 

Water Pollution Conml Federation. Wrahington. D.C. 

Welch, P.S. 1952 Imnology. M&nw-Hill Book CO., New York. 

DEMETON 

REFERENCES CITED 

Butler, PA. 1964. COmwrd.l fiihery isvestigatbm. P.ges 61B in Pedicide vildliie 

atudiea. 1969. U.S. bh md Wildlife Wee Cinuh 149. WMhingtOn, D.C. 

fuh. Tnnh Amer. Fish Sa 67: 39. 

Ludenunn, D., rad H. Neurmnn ube die wirh~ng der dtlichen konwdnsedd. 

tidde auf die tiere de auasr.mers.'higu fur schdlingskunde und PILuneMut~. 

s:5. 

Mek. €3.. md W.A. MeAllkter. 1970. lnseetidde ausoeptibility of *we mmmon fuh 

funily repesentativea. Tnna Ama. hh. Sa 992l. 

toxic levels of patieides Tnna Amer. Fiah k 101: 817. 

-of oumw~terfiishesTnns.Amer.Fiah.~81:175. 

Henderson. C.. md Q.H. Mering. EX8. Toxicity of organic pimphaw i- to 

YcCmn, JA., md RL Jq. WR Vcrrebnl d.nuge to Mueeilb expxed to rutely 

Pickering, Q.H., et d. 196L Tbe toxieity of ap.nie phoaphow inaect*idpa to different 
Sanders, H.O. 1912 Toxicity of come iMeetiddes to four SpeCKa of malamstracan 

wtxmns. Burnu of Sport Fisheries and Wildlife, Tech Plpu No. 66 US. 

Department of tbe Inkria, Washington. D.C. pp. S-19. 

We&+ C.M. 1968 Thl Qtamirution of cholinedteme in the hain timueof thR0 lpgies 

of fnshwstu Tub md ib iructividon in vim. Emlogy, 89: 194 

We&. C.M. 1968. Reapo~ of fish to sublethal expos- of -* phmphom 

iweeticidea. Sew. and Ind. Wastes. 31: 580. 

Web. C.M. 1961. PhVsologieal effect of organic phosphorus inseetieides on several 

speciesol fish. Trans. Amer. M. Soc. 90: 143 

Web. C.M.. and J.H. Gakstatter. 1964s. Detection of -ticides in water by biwhemical 

aaaay. Jour. Water Poll. Cant. Fed. 36: 240. 

Weisa. C.M.. and J.H. Gakatatter. 1Wb. The decay of anticholineste- activity of 

organic phmphorua inseeticidea on storage in waters of different pH. Advances in 

Water Poll. Research, I:&%. 

GASSES, TOTAL DISSOLVED 

REFERENCES CITED 

buck. G.R, et al. 1975. Monality. saltwater adaptation and reproduction of fMh expeed 

to gas supmaturated water. US. Environmental Roteaion Agency, Western Fish 

Toxicology Station. Cawallis. OR. Unpublished reprt. 

Clay, A.. et al. 1975. Erperimental induction of gas bubble W in menhaden. 

PRsented at the American Fisheries Society, September 1975, Ira Vegu, Nev. New 

England Aquarium, Boston. 

Dewley. EM., and WJ. &I. 1975. Effect, of various mnoenhLions of diaaolved 

atmmphenc ge on juvenile chinook salmon. Onmrhynchw tdmqfda, and ateelhesd 

trout. Fish. Bull. 

Dawley, E, et al. 1975. BioaaMp of total dissolved gas pmu~ National Mviw 

Fisheries Sen&, Seattle, Waab Unpublhhed ~prt 

DeMont, J.D., and RB'. Miller. 1972 first reported incidewe of p bubble dkw in the 

heated effluent of a steam electric generating station. Roe. Bth land meeting, 

Southerat Assoe. of Game and Fiah Commi\aown 

Ebel, WJ., et al. 1975. Effect of atmospheric gas superssturation carved by dams on 

salmon and steelhead trout of the Snake and Columbia Riven. Final Report. Nhwest 

Fisheries Center. National Mpriw Fisheries Service. Seattle. Wash. 

Keup.L.E 1975. How toreadafishkill. WaterandSew%ge Work8,121:18. 

Lindmth, A 1957. Abiogenic gas iupnstmtion of river water. Arch. Iiydrobiology, 59: 

589. 

Mrlouf, R 1972 Oenurenoe of g.s bubble disease in W species of bivalve mollusks 

Jour.Fish.Res.Bd.Can.B:SBB. 

Mlsoello, RA, et al. 1975. Evaluation of alternative mlutio~ to gas bubble disease 

monality of menhaden at Pilgrim Nudear Power Station. Yankee Atomic Electric Ca.. 

Weatborn. ~~~ ~~ ,~~~ M- ~ YAEX-IoB7. ~~~ 

Kaoonal Academy of Saenapa. National Academy of hgimennp. 1974 Water quality 

cntena. 1972 V.S. Government hung Office. Washington. D.C. 

Nebeker. AV., et al. 1975. Effects of gas supmaturate -water on freshwater inverteb 

ntes. Roe. Gas Bubble M Worlrshop. Battelle Northwest, Energy Reeeych lad 

Development Agency Spgial Report 

Nebeker. AV., et al. 1976~. Nitrogen, oxygen, and carton dioxide as factom dfdng fMh 

BhVd h 8y Npenatmted water. h. her. Fish. SOe 

Nebeker. AV., et al. 197%. SvVival of mho and scckeye salmon smolt, in seawater dter 

exposure to gas 8upenaUMt.d water. TraM. Amer. Fish Soe. 

Renfm, W.C. 1W. Gas bubble morlality of flahes in Gilveston Bay, Tex. h. Amer. 

Fish. 5% 92320. 

Ruder. RR 1974. Gas bubble disease: Monalities of mho salmon, Chmrh&ud &A, 

in water with mustant total gas presure and different oxygen-nitrogen ntim. 

National Oceanic and Atmm. Admix, Natl. bhr. Fish. SW., Northwest Fish Center, 

Seattle.WMh Unpublished muscript 

Rulifmn, RL and G: Abel. 197L Nivogen supenaturation in the Columbia and Snake 

Riven. Ted Rep; TS-Cr%'/O-Z%0162 U.S. EnvLonmental Rotenion Agency, 

won X, Seattle, Wash. 

Van Slyke. D.D., et al. IW. Studies of gaa md electrolyte equilibri. in Moo& XVIII. 

Solubility and phyxieal state of atmpheric ni-n in blood fell8 and plrama Jour. 

Biol. Chem. 1&:571. 
GUTHION 

REFERENCES CITED 

Adelmn. I.R, md LL Smith. Unpublished data Deprtment of htomlogy, Fisberiea 

and W'ildlife, University of Minnesota, St Paul. 

Benlre, G.M.. md S.D. Murphy. 1974. Anticbollnesterase setion of wthyl purtbjon, 

ppnrhion and ezinphoamethyl in mi@ and fmh: Onset ud remveryof inhibition. Bull. 

Envimn. Contam. Toxiwl. 12: 117. 

Butler, PA 1963. CommerCial fuheries investiptiona. In Peaticirk-wildlife rtudim 

during 1961 and 19EZ US. Fish and Wildlife Service Circ 167. W.abin%on, D.C. 

Caner, F.L 1971. In tIw studies of brain acetylcholines- inhibition by wkwpbosp 

bate and earbarnate inseetieides in fish PhD. Dhenation, LouLiuu State Univerrity. 

Baton Rouge 

Coppage. D.L 1972 Orgmophosphate perticidea: Speific level of bnin ACHE inhibition 

relawd to denth in sheephead minnows. Trans. Amer. Fish Soe. 101:534 

Coppcge, D.L. and T.W. Duke. 1971. Effeets of penicidfa in eatuvica dong the gulf and 

mutheart Atlantic mas-. In C.H. Schmidt, ed. F?a+nga of the 2nd gulf mt 

mnference on mquita suppmion and wildlife management National Mosquito 

Control-Fish and Widlife Management Coordinating Committee, Wuhingtan. D.C. 

Coppage, D.L and E Mrtthews. lPi4. Sbon term effectsof orgawpholph.te pticidca on eholinfaterace of estuarine fuhea and pink ahrimp. Bull. Envimn. Contam Toxiool. 

11: 483. 

Coppage. D.L et al. In preaa. Brain seetylcholinesterase inhibition in fmh M a diagnmk 

of environmental poisoning by malathion. 0,Mmethyl S(l.24imrbethoxyethyl) 

phoaphorodithioate. Pesticide Biochemistry and Physiology. 

Davis, H.C.. and H. Hidu. 1969. Effects of pesticides onembryonic development of dm 

and optera and on survival and growth of the larvae. U.S. Fish and Wildlife Service, 

Fishery Bulletin, 67: 993. 

aton. J.G. 1970. Chmnic malathion toxicity to the blvegiU (kpomis &inrs 

ILfinesque). Water Researrh,4:6iS. 

Flinl, D.R. et d. 1970. Soil runoff, leaching. md .dsorption .nd water stability studies 

with guthion. Chew Rept No. m. 

Gaufn, A.R. et d. 1965. The toxicity of 10 organic insecticides to,vviow 4ustic 

invertebratea. Water and Sew. Worlw. 112276. 

Hollnnd. H.T., et al. 1967. Use of fmh brain acetylehalireterme to monitor pollution by 

organopbmphoruc pesticides. Bull. Envimn. Contan Toximl. 2: 156. 

Jensen. LD., and AR Caufn. 1966. Acute and lpng term effeets of organic inmaicides on two Species of stowfly nriadr. Jour. Water PolL Cont Fed. 98: 1zTJ. 

btr. M. 1961. Acute tadcity of mme organic inseeriddes to thrse @ of vlrnonida 

and to the thresapine sdddeback. Tm. Amer. Fish. Soc 90:!264. 

Wv. Y., and S. Sai+g. 1989. Se~itivity of pond fuh to mtnion (uinphmethyl) ud 

panthion. Bamidgeh.21:67. 

Macek, KJ.. et a]. 1969. The effects of temperature on thesuseoptibilityofthe bluegills 

nnd rainbow mut to selected pesticide% Bull. hvimn. Conram Toxiad. 3174. 

Macel, KJ., and W.A. MeAllister. 1970. Insecticide susoeptibility of some mmmon fmh 

family repmntativea. TMa her. Fish. sOe.99:P. 

Meyer. F.P. 1965. The experimental u8e of guthion as a selective fvh eradicator. Trms 

Am. Fiah Sor 44:'ZX. 

Nebeker, A.V.. and A.R Gnuh 1964. Bioaarays to determine psticide todcity to ilw 

amphipal mwn. Gummanu hcwtris. Roe. Utah Ad. sd. Ma and Letters. 

41:64. 

Piekering, Q.H., et al. 1962 The toxicity of organic phaphonu insecticidca to different 

species of wmwater fmhea Trana. d. Fish. k 91:1'75. 

Ponman. J. 1912 Resulfs of acute testa with marine organisms. usingitandvd methods. 

P.ges 212-217 in Ruivo. ed. Mene pollution and se~ life. Fishing Nem (Books) Ltd.. 

London. 

PmL G.. and RA. Lersure. 1974. Sublethal effect of malathion to three dmnid spied. 

Bull. hvimn. Conulm. Toximl. 12: 912 

sudem. H.O. 1969. Toxicity of pesticides to the matacean. Gammanu hcwtri~ BWU 

of Sport Fisheries and Wildlife Twb. Paper No. 25, US. Depanment of the Interior, 

WMbinRWn. D.C. 

Wen. 

cn~- 

KO. 

US. 

197!2 

Deprtment 

Toxicity of 

of 

wome 

the Interior. 

i&dea 

Bureru 

to 

of 

four 

Sport 

apeEiea 

Fisberiea 

of -

and Wildlife 

- 

Tecb Parer No. 66. Wnhington, D.C. 

Wen, E.O., ud O.B. Cape. 19BB. Tne relative toxicities of wed ptiddea to nrids 

Tucker. RK. md D.G. Crabtree. 1970. Handbook of toxicity of peatiades to wildlife. 

Bvreru of Spwr FLberica ud Wildlife Raavor F'ubl. No. sl US Ikprrmantof tbe 

Interior, W.rhin%on, D.C. 

We% C.Y 1868 Tbe determimtion of eholinertanse in the bnin tissue of throe spedea 

of freshwater fuh ud itr rrivation is viw. Emloay. S: 1% 

Weim, C.Y. 1868. Rasponr of ri to mbleth.l expama of 0rp.niC 

id& Sew. and Ind W.stea, 31: 580. 

W&. C.M. 1961. Physiological effects of wp.nic pbaspkru ineriddps on severr.! 

Web, CX, lad J.H. c.Lsuttu. 1864 Detection of mdea in Gtu by bkkw 

ofLhree~ofatoneiliea~ol.~oceanog.l3:1lz 

.Fe&of Ti TMa Amu. Fhh kso: 143. 

UUY. Jw. Water Poll. Cont Fed. 86: 240. 

a 
HARDNESS 

a 

REFERENCES CITED 

Doudomff, P. and M. btz. 1953. Critical review cf li+rstweon the toxicity of induarrial 

wastes and their mmponenh to fish. 11: The metabasa.d@.Sew.andind. Wasted,25: 

Bm 

National Academy of Science. National Academy of Engineering. 1974. Water quality 

criteria. 1912. US Government tinting Office. Washingon, D.C. 

National Technical Ad-r). Committee to the Secretary of the Interior. 1963. Water 

quality enterin. US. Government tinting Office. Washington, D.C. 

Sawyer. C.H. 1960. Chemisq for sanitary engineen. McGraw-Hill Book Co.. New York. 

Stiff, MJ. 1971. Copper/biearbonatP equilibria in mlutio~ of bicarbonate ion at 

mncentrations similar to thm found in natd water. Water Research. 5: 171. 

MALATHION 

REFERENCES CITED 

Andemon. B.G. 1W. The toxicity of organic insecticides to hphnia Semad Seminar on Biol. Pmblems in Water Pollution. Robert A. Taft Sanitaryhgineering Center Tgh. 

Report W603, Cincinnati, Ohio. 

Bender, M.E 19%. The toxicity of the hydrolysis and breakdown pduaa of maLthion to 

the fathead minnow (Anuphaler pmmcloa Rainesque). Water &a 8:571. 

Bwnu of Sport Fisheries and Wildlife. 1970. Resourn Publication 106. W.abiagton, D.C. 

Butler, PA. 1963. Commercial fisherie investigations. In Pd& wildlife studies 

during 1961 and 1962 U.S. Fish and Wildlife senice cirr 167. Wdiagton, D.C. 

Coppage, D.L, and T.W. Duke. 19il.~fect3cf~ticidesinestu~es~ongtbe~~ 

southeast Atlantic masts. In C.H. Schmidt, ed. Roceeding~ of the 2nd gulf mt 

mnference on mosquito suppression and wildlife mamrnent. Natiod Mosquito 

Control-Fish and Wildlife Manapmeat Ccardimting Committee, Wdngton. D.C. 

Coppage. D.L. and E. Matthema. 1974. Short term effects of organophosphate peatieides on cholinestem of estuarine fiahes and pink shrimp. Bull. Envirun Coatam Toxiad. 

11:w. 

Coppa@?, D.L. el al. 1975. Brain acetylcholinesterase inhibition in fuh M adiagnosis of 

envimnmental poisoning by malathion, 0, Odirnethyl s(l.Zdiearbetboxysthy1) 

phmphorcdithioste. Peaticide Bioehemistrj and Physiology. 

Cope. O.B. 1965. Sport fuher). investigations. In The effect3 of peaticides on fiah and 

wiidlife. Fiah and Wildlife Service Cidar 266, US. Depanment of the Interior, 

Washinkon. D.C. 

Da\% H.C.. and H. Hidu. 1989. Effeets of pesticides on embryonic development of dam- and optera and on sUrnval and gmalh of the larvae. U.S. Fish e,nd Wildlife Service, 

Fisbery Bulletin. 6i: 399. 

Eaton. J.G. 1970. Chronic malathion toxicity to the bluegill (kpomis moFTochltlyl 

Rafinesque). Water Res. 4:6?3. 

Eiehelberger, J.W.. and JJ. Licbtenterg, 1971. Persistence of peaticidea in river water. 

hvimn. sd. Technol. 5541. 

Eider, R 1988. Acute toxiatiea of idades to marine dapd Crust.~1- aa 16:m 

Eisler. R 1970. Acute toxicities of organochlorine and organophosphorus i~eetiddea to 

atuvine fiahes. U.S. Burerru of Sport Fisheries and Wildlife, Tech. Paper 46, 

Washiwn. D.C. 

Henderson. C.. and Q.H. Pickering. 1958. Toxicity of organic phosphorus insecticides to 

kh. Tm. Amer. Fish Sor S: 39. 

M, RE. and WJ. Collins. 1974. The susceptibility to sekcted insectidea md 

umtylcholiwsterase activity in a laboratory mlony of midge larvae, Chimnmnus 

trmonr (Dipters: chimnomidae). Bull. Envimn. Contarn. Toximl. E62 

Kaq M. 1961. Acute toxicity of some organic insecticides to three species of Sllmonids 

and to tbe threespine stickleback. Trana. Amer. Fish Soe. W.%. 

Mad. KJ.. and W.A. McAllister. 1970. Inseeticides suaeeptibility of ~omemmmon fmh 

family repRsentatives. Trans. her. Fish. Soe. 99:ZJ. 

Mount, D.T.. and C.E. Stephan. 196i. A method for establishing sxeptahle toxiant limits 

for fish-dathion and the butoxyethanol ester of 44-D. Tram Arner. Fish Sa 21: 

185. 

Mulla. Y.S., and A.M. Khssauinah. 1969. Laboratory and field evaluation8 of Iddea 

.gainst ehimnomid midp. Jour. Ecan. Entomol. 6237. 

Puis. D.F.. et al. 1975. Rates of depadation of malathion by Ma klafed fmm 

aquatic system. Envimn. Sci. Technol. 9: la. 
Piekering. Q.H.. et d. 1962 The toxicity of orpanic phmphom insectickb to different 

spies of wmwater fishes. Trans. Amer. Fish. SOC. 91:175. 

pmt, G., and T.R Sehroder. IKI. The toxicity of four iruecticides to four dmonid 

species Bull. Envimn. Contam. Toximl. 6: 144. 

Sanden, H.O. 1W. Toxicity of Wticides to the crustacean. Cammnu WnS. BmU 

of Sport Fisher% and Wildlie Tech. Paper No. 25. U.S. Department of the Interior. 

WashingWn. D.C. 

Sanden. H.O. I972 Toxicity of some insecticides to four vies of mdrmsWn 

cmsu~oedns. Bureau of Sport Fisherier and Wildlife Tech. Paper No. 66, US. 

Deportment of the Interior. Washiwpn, D.C. Sandera. H.O., and O.B. Cope. 1966. Toxicities of several peatieides to two Of 

cladmeram Tram. Amer. Fish. Soc. %:I&. 

Sanders, H.O.. and O.B. Gap. 196% The relativetoxicitiesof several peatiadesto lyiads 

of rhree species of stoneflier. Limnol. Oeeanog. 13: 112 

Weiss, C.M.. and J.H. Gakntatter. 1964. The decay of anticholinesteRas Vtinty of orpanic 

nhaaohom insecticida on storag~ in waten of diiferent pH. Advm in Water 

k'oll;tion Fk%e.rch, 1:83 

YIGANESE 

REFERENCES CITED 

F.irbndge. RW.. ed 1966. The encyclopedia of axanogmpby. Reinhold. New York. 

Griffin. A.E 1960. Si&icanoe ud ~moval ofrmnganeee in water aupdica. Jw. Amer. .. Water Worka Asan. 52: 1926. 

Amer. Water WorLs hn. 52: W7. 

Illig. G.L Jr. 1960. Use of durn heurnetaphosphate in man~.neest.biliutioa Jw, 

McKee. J.E.. and H.W. Wolf. 1963. Water quality eriterir State Water Quality Control 

Natioml Academy of Sciences, National Aeademy of Engineering. 1974. Water gdity 

Riddirk, J.M.. et d. 1958. Iron and manganese in water suppliea Jour. Amer. Water 

Board. Sacramento. calif. Pub. 3-A. 

criteria, 1972. US. Government Rinting Office. Washinkon, D.C. 

Works Asan. 54: 688. 

sollman. T.H. 1457. A manual of pharmacology. W.B. Saunden, Philadelphia 

METHOXYCHMR 

REFERENCES CITED 

Bdmer. LE.. and D.R Nimmo. 1974. Methods to ae.s&as effeets of combinations of 

tohtr, salinity, ud temperature on wtuuine minula Roe. 9th Annd Conference on Trace Submama in hvironmenul Health, University of Yiamuri. Columbia 

Butler. PA 19n. Influence of petiadea on mMne emsyatems. Roe. Royal Soi bhn, 

Eisele. PJ. 1974. The effeets of mqbxyehlor on quatic invertohate papuktiona md 

communities. PkD. %is, University of Michigan, Ann hrbor. 

Heath. RG., et d. 1972 Comparative diewry toxicities of peStieidea to birds. Burerruof 

Spn FLberiea 4 Wildlife. Widlife Report No. 1% US. Depvrmeat of tbe Intcrior, 

Washinrton D.C. 

in: sn. 

Hodge, Hk., et d. 1954. Short term oral toxiaty testa of methoxyehlor Q ntrd +. 

Jour. Pbumaml. Exo. , Theraav -,.-- 99. la 

Kern. S.. and R Earnest 1974. Acute toxicity of a, innecrjddm ta dipad hr Mamtilu. Womb Fish ud Game, 60: B 

Lehnun.AJ. 1W. Summvlesof peaticidetodaty.&socofFoodudhvgOffdrbof 

the US.. Topeka h. pp. 14. 

Ym, J.W.. ud PJ. ble. lgls. The effem of methoxyeblor on 4ucie biota U.S. 

hvimnmenul Rotenion Agency, Ecologid Re. Senea, So. EPA-WXun& U.S. 

Covemment Printing Office. Wpahiunpn, D.C. 

, -,.-- Kern. S.. and R Earnest 1974. Acute toxicity of a, innecrjddm ta dipad hr Mamtilu. Womb Fish ud Game, 60: B 

Lehnun. AJ. 1W. Summvles of pesticidetodaty.&socofFoodudDrupOffdrbof .--- -...r-_._ rr ~ - 

Ym, J W.. ud PJ ble. lgls The effem of methoxyeblor on 4utle biota us 

hvvonmenul Rotenion Agency, EC0logx.i Re. Senea, So EPA-WXun& US 

Covemment hung Office. Wpahnpn. D C 
McM, RL. et d. 1911. Biodegradable analog3 of DM. Bull. World Health Omanhsmden. H.O. 1972 Toxicity oS sue iwetieidea to four sped= of daooanCan 

BURIU of spwl Fiahenes and Wildlife Tech. Paper No. 66, US. 

Stkkd, LF. 1919. Pesticide residues in bin$ and mammals. Page4 254-312 in C.A. 

tiW.44m. 

Governaclrt Rinting Off=. Washinkon. D.C. 

Ed& ed. Environmental pollution by peaticides. Plenum Reas. New York. 

HIREX 

REFERENCES CITED 

Borthwick, P.W., et al. 1973. Accumulation and movement of mirex inselwted esturka 

of South Carolina, 196s-71. Pmtiaded Monit Jour. 1:6. 

Bookhout. C.G.. et d. 1972 Effm of mirex on the lwal development of two enbs. 

Water,Air.and Sail Pollution, 1:Ia. 

Heath, RG., and J.W. Spann. 1973. Pesticidfs in the environment: A mntinuing 

mntmversy. lntemont Med. Bmk C~Q., New Yark, pp. 01-4s 

Heath. RG.. et al. 1972 Comparative dietary toxicities of pesticides to bi&. Bmu of 

Spn Fisherie and Wildlife, WildliIe Repan KO. 152, US. Department of Inrerior. 

Wahington. D.C. 

Hydf. K.M., et al. 1974. The effect of mirex onchannel caUishpduction.Trana. Amer. 

Fish. Soe. 1(n:366. 

Kaiser, K.L.E. 1974. Mirex: An unremp& conraminant of fihes from Me Ontario. 

Seience 185:523 

Lowe. J.I.. et 81. 1971. Effwts of mirex on selected estuarine organisms. Pap 171-186 in 

Ludke. J.L., et a]. 1971. Toxicity of mirex to crayfish. Ifomdm blandinqi Bull. 

Naq\i. S.M., and A. de Ih Cw. 1973 Mirex inmrporation in theenvlonment: Residue in 

Stickel. WH., et al. 1973. Pesticides and the envimnment: A mntinuing mntmvmy. 

Tapau. M.E., et al. (19%). Seasonal effecfs of leached &x on selected esuurine 

Van Valin, C.. et al. 1s. Some effects of mirex on two wm water fihes. Tram. Amer. 

Wolfe. J.L., and B.R. Nonnent I973 Ammulation of mirex midues in selected 

organkms after an aerial treatment. Mississippi, 1971-72 Pesticides Monit. Jour. 7: lrz 

Transactions of the 36th h’onh American Wildlife Wum Conference. 

Environ. Conram. Toxicol. 6:89. 

non-tar@ organisms, 1972 Pesticides MoNt Joui. 7: 104. 

lntermnt. Med. Book Corp.. New York, pp. 43747. 

animals. hrch. Envimn. Contam. Toxicol. 3:371. 

Fish. Sa. 9i: 1%. 

NITRATES, NITRITES 

REFERENCES CITED 

Gillette, U, et al. 1952 Apprakl of a chemical we problem by fiah toxicity tests. 

Humemn. RH.. et d. 1971. The nitrate mtuation in Illinoia Jour. Amer. Water Works 

Sewage Ind Wastes. 24: IS. 

bn. b3:m. 

Klinger, K 1957. Sodium nitrite. a slow acting fih poison. Schweu 2 Bydml. 19(2):565. 

Kwpp. G.L, and G.F. Arb. 1973. Ammonia toxicity levels .ad nitrate tolennce of 

channeles~~hTheProgRsJiveFLahCulturist,35:221. 

U&y, EF. 1972 Role of bacteria in the nitrogen cycle in Inkes. Water Pollution Conml 

Reaeuch Series, (EP 210:16010 EHR WE), US. hvimnmenul PmWon Agency, 

US. Government Rinting Office, Washington. D. C. 

National Ademy of Sciences. 1972 Aocumulation of nitrate. National Academy of 

Sdenw, Wlshington. D. C. 

Narionsl Ademy of Sciences. National Academy of Engiarering. W97C Water quality 

criteria, 1972 US Government Rinting Offioe, Wrshingmn, D. C. 

Public Hdth Sernce. 1961. Groundwater mnumination: prmedinga of 1961 symposium Tph. Rpt W61-5, RA Taft Sanitary Engineering Center, U.S. Public Health 

Senice, Department of Health, Education and Welfare, Cincinnati. Ohio, 
Ruso, RC.. et d. 1974. Acute toxicity of nitrite to rainbow hut (Sdmo pidmi). Jour. 

Fish. Res. Bd. Can. 31:1663. 

Rueo, RC.; and RV. Thurston. 1975. Acute toxicity of nitrite tocutthroat tmut (Sdmo 

elark'). Fkherka Bicamy Lsboretory Tech. Repn No. 754. Yontms State UwvvaiW. 

tr. 

A. 1965. Studia on fsh culture in filtered doeed-drculating quaria 11. On the 

arp culture exprimenu in the syayatems. Bull. Jap. Scc. sd. hab. 31:916. 

httelmarher, P.G. 1962 Methemoglobinemia fmm nitrate. in drinking ntu. SehrReihe. 

Ver. W.aser-.Boden-u. Lufthyg. Ko.21. *her, Stuttgan 

Smith, C.E. and W.G. Williams. 1974. Erpenmntal nitrite toxicity in rainbow hut and 

chinmk dmon. Trans. Amer. Fish. See. laS: 389. 

Stewut. B.A. et al. 1967. Nitrate and other pollutants under fields and feedlots. Envir 

Sn. Tech. .~~~~ I:%. ~ .~ 

Trarm. F.B. 19%. The .cute toxicity of mm mmmon salts of mdium. pomsium. and 

ddum 10 the mmmn bluegill (Lrpnla marrochinu Railwsque). Ra Ad Nat 

Sci.. Philsdelphie 106: 185. 

Vipil, J.. et d. 1965. Nitrates in municipal water supplies muse metbemoglobinemia in 

infanta. Public Health Rept 80: 1119. 

Wdlen. I.E. et d. 1957. Toxicity to Gambuaio affinisof certain purechemiakin turbid 

watem. Seam Ind. Wmm. 29: 695 

Wdton. G. 161. Survey of literature relating to infant methemogiobinemi. due to 

WVestin. D.T. 1974. Nitrate and nitrite toxicity to salmonid fuhes. The wve Fishnitraw-cvntaminated water. Amer. Jou. Public Health. 41:M. 

C"lt"rist. %!@& 

WoUf. LA, and Wakrman. 1972 Nitrates, nitrite., and niuoaamine. Science, 177: 15. 

OIL AND GREASE 

REFERENCES CITED 

Bellan. et .I. 1972 The sublethal effeeta of a detergent on the reprodudion. development, 

and settlement in tbe polyeh.etous annelid Gpitclb mpilda. Yyim BWlo([y. 14: uE3. 

Blumer, M. 1970. Oil matahnation and the living reanvces of the .e~ Fmd ud 

Ag~iculture Orgsnizdtjon Tech Cod. Rome. FIR: MPrlO/Rl,llP. 

Bugbee, S. L. ad C. Y. Walter. 1819. The ~spow of maaminvertebratea to penline 

pollution in a mountain stream. Page 7% in Revention unl mntml of oil spilla, 

proeeedinga of symposium .. March lb17. W.ahiMton, - D. C. 

Diaz-Pilerrer. 1962 The effects of an oil spill on the ahore of Cuania, Puerb Rim 

Hampaon. G. R,. and H. L Sanden. 1989. Local oil spill. Oceanus. 15:8. 

Imbmn. S. M.. and D. B. Boylan. 19TJ. Mfect of seawater mluble fraction of kemeene on chernotaxis in a marine snail, Naamrius ddclua. Nature, 241: 213. 

Johns Hookins UniveniQ. 1956. Final report to the water quality submmmittee of .the 

(ahtract). Ass. Isl. Mar. Late, 4th Mtg. Cum, 12-13. 

Am&& Petmleum I&titute,Project~49.41. 

MeKee and Wolf. 19'3. Water quality m'teria Slate Water Resource Control Board. 

Saeramento. Calif. Pub. No. %-A. 

Yimnov, O.C. 1961. Mfeets of low mnaentrations of ptroleum and its pmducts on the 

development of me of the Black Sea flatfish. Vop Ikhtiol. 7: 557. 

Yimnov. 0. G. 1970. The effm of oil pollution on flora and fauna of the Black Sea In 

Rmedirq: FA0 mnferenoe on marine pollution and its effects on living resouroes 

and fish. Rome. December 1970, 692 Food and Agriculture Organization of the 

United Natiom. 

Moore. S. F.. et d. 1973 A pmliminwy assessment of the environmental vulnerability of 

Machi= Bay. Maine tooil superlankem. Report No. MITSG 734. 

Nelsan-Smith. A. 1971. Effects of oil.on marine 0lmt8 and anid. Paae~min P. 

Hepple. ed. Water pollutlon by oil. London. . . 

Nelmn-Smith, A. 1973 Oil pollution and marine emlogy Plenum PreSB. New Yo&. 
PARATHION 

REFERENCES CITED 

Bill&, R. and deKinkelin. 1970. Sterilization of the Mticles of guppies by me8116 of non- lethal doeeaof parathion. Annales D Hydrobiologie, 1(1):91., 

Burke. W.D., and D. E Ferguson. 1989. Toxicities of four imticides to &stant and 

susoeptible mosquitofuh in static and flowingmlutions. Mquito News.291):96. 

Carter, F. L 1971. In mw studies of brain aoetylcholine~terme inhibition by organophmp 

bate and earbarnate iaseeticides in fish. Ph.D.&nation, &MUM State University. 

Baton Rouge. 

Coppage, D.L 1972 Orghnophcephate peaticid-: spedfic level of brain AChE inhibition 

related to death in ~heepshead minnows. TRM. Amer. Fish. h.'l01:594. 

~oppage, D.L. and 1. W. Duke. 1971. Effecta of pesticides in estuaries along the gulf and 

mutheas Atlantic mta. In C.H. Sehmidt, ed. Roeeglinga of the 2nd Gull Caa 

Conferem on Mcaquito Suppression and Wildlife Management National Mmquitn 

Conuol - Pish and Wildlife Management Cmdinating Committee. R'sahingvn, D.C. 

Coppape. D. L. and E. Matthews. 1974. Short term effead of organophaephate pesticides 

on cholinestenues of estuarine fuhes and pink shrimp. Bull. Environ. Contam Toximl. 

11:w. 

Coppage. D. L, et al. 1975. Brain scetylcholinestenue inhibition in fuh a8 a di.gnoaie of 

environmental poisoning by malathion, 0.O-dimethyl S(1,ldicarbethoxysthyI) phob 

phorodithioate. Pesticide biochemisuy and phpiology. 

Dowden, B.F. 19%. Effecta of five insecticides on the oxygen mmumption of the bluegill 

sunfish. LqmnW mnnahinu. Ph.D. Thais. Louisiana State University. Baton m. 

Eaton, J. G. 1970. Chronic malathion toxicity to the bluegill, (Lcpmna mannhirw 

Winesque). Water Reseueh. 4:673. 

Eichelberger. J.W., and J. J. Lichtenberg. 1971. Penistenoe of pesticided in river water. 

Environ. Sci. and Technol. 5541. 

Enrimnmental htection Agency, Office of Pesticides Program. 1975. Initial sientific 

and miniemnomic review of parathion. Repon. No. EF'A-540/1-75ml. US. Environ- mental htgtion Agency. Natiod Technical Information service, Springfield, Va 

Gaufin. A. R. et al. 1465. Toxicity of 10 organic insectjades to various aquatic 

inuertebnrter. Water and Sew. Works, 112276. 

Gibmn. J. R. et al. 1969. Sources of ermr in the use of fish-brain aoetylcho1ioeste-k 

activity as a monitor for pollution. Bull. Envimn. Contam. Toximl. 437. 

Jeaaen, L D., and A. R Gaufin. 1964. Lanp term effects of organic idcides on two 

species of stonefly naiads. Trans. Amer. Fish. Soe. 93: 357. 

Korn. S.. and R. Earneat 1974. Acute toxicity of 20 insecticides to striped bass. Mm 

&ilk. Califomin Fish and Game, 60: 128. 

Lahay, M.. and S. Sarig. 1989. Sensitivity of pond fuh to mtnion (azinphwmethyl) and 

parathion. Bamidgeh. Bull. Fbh. Cult Israel. 267. 

Leland. H. V. !I. 1968. Bimhemid facton affecting toxicity of parathion and selected 

lrnalop to fuhes. Univemity of Michigan, Ann Arbor. 

Lowe. J. I., et al. 1970. Laboratory bioausys. In Fhgrss report for f4 yy 1969. 

Pesticide Field Station, Gulf Breeze, Fla U. S. Fish and Wildlife Service Cue F. 

Ludke. J. L 1970. Mechanism of resistance to parathion in mmquitofuh. GambvM 

afini8. Ph.D. Thesis, Mississippi State Univenity. Univemity. 

Mount, D. 1.. and H. W. Boyle. 1969. Parathion - y9e of blmd mnoentmtion to diagnose 

mortality of fmh. Environ. Sci. and Technol. 3: 1163. 

Mulla, M. S.. and A M. Khasa&nah. 1969. Lsborntory and field evaluation of larvicides 

againat Chimnomic mid-. Jour. Emn. Entomol. 62:s 

Murphy. S. D.. et al. 1963 Comparative anticholineaterase action of organophosphorus 

inseeticides in invertebra-. Toximl. Appl. Phannaeol. 12:P 

Pcst. G.. and R A Leasure. 1974. Sublethal effect of malathion to thRe dmonid sped-. 

Bull. Envimn Cordam. Toximl. 12: 312 

Sanden. H. 0. 1972 Toxicity of some insecticides to four species of malamstracan 

mwtaeeans. Bureau of Sport fisheriea and Wildlife, Tech. Paper No. 66. U. S. 

Depanment of the Interior, Wsahingvn. D.C. 

Wen. H. 0.. and 0. B. Cop. 19%. Toxicities of several pesticides to two species of 

CIadooenuls.TRlu. her. Fish. Sa 95:165. 

Spacie, A. 1975. Acute and chronic parathion toxicity to fuh and invertebrates. U. S. 

Environmental Pmteetion Agency Emlogical Reeeareh Series. 

Web, C. M. 195%. The determination of cholinestersse in the brain tissue of three a+es 

of freshwaterfhhandiuinaeti~.atiDninvivo.Eeology,39: 194. 

We&, C. M. 1959. hponse of fish !A sublethal expmures of organic phoaphom 

insefticides. Sew. and Ind. Was@, 91: 580. 

Web. C. M. 1961. Physiological effect of organic phmphorus ilueeticide on severd 

species of fuh. Trans. Amer. Fish. Soe. 90: 143. 

We*, C. M., and J. H. Gakstatter. 1964. The decsy of anticholinestersse activity of 

organic phcephom insecticides on storage in waten of different pH. Adan- in 

Water Poll. Res. I:@. 

0 
. 

REFERENCES CITED 

Bell, H.L 1971. Effect of low pH on the survival and emergem of equatic insetu. Water 

ILea.5:313 ~~~~ ~ ~ 

Butterfield. C.T. 1M. Bacterial propetties of free and mmbined available chlorine. Jw. 

Am. Water Works Assn. 40; 1%. 

Gpuro. LRA. 1970. Oeeawgrnphy for practicing engineers. Barnes md Noble, Inr, 

New York. 

DeUino. JJ.. and G.F. Lee. 1971. Variation of manganese. diamlved oxygen and relaled 

chemical parameters in the bottom waters of Me Yendota, WM. Water b. 5 lXY/. 

Ewpem Inland Fiaheriea Advimry Cornmisaion. 1969. Water qdity aiteria for 

European freshwater fish-extreme pH values and inland fuheries. ,Repred by 

EIFAC Working Pnrty on Water Qdity Criteria for Eumpean hbwbkr hh. 

Water Reseuch. 3: 593. 

Joned. J.RE. 1964. Fish and river pollution. Butterworth, London. 

Lngelier. W.F. 1936. The rJlalytical mnml of antiamion water tRltment J&. 

Water Worka Asan. 28: 1500. 

Mnunc ~.. D.I. 

~ .~ 1m. ~ 

~hmnic effect of low PH on fa- minnow mm+v.l, rod and 

reproduction. Water Rea. 7:W 

uiterin. 1972 US. Government Printing Office. Washin@n. D.C. 

Enviro. Engineering Div., ASCE 100: No EEL Ror Papers 1W43.539. 

National Aodemy of Soenaes, National Academy of Engineering, 1974. Water qdity 

Reid, L.C., and D.A. Cprlson. 1‘374. Chlorine disinfection of low temperature watem. Jour. 

Sawyz, C.N. 1960. Chemistry for saniuus. engineers. MeGraw-Hill Book Co.. New York. 

Stumm. W., and JJ. Morgan. 1970. Aquatic chemiaty. John WIley and Sons, I=. New 

Zirino. A., and S. Yamamoto. 1572 A pHdependent model for the chemical sppi.6~0 of 

.. 

York. 

mpper, nnc. cadmium and lead in seawater. Limn. and oeploog. 11: 661. 

PHOSPHORUS 

REFERENCES CITED 

Fletcher. G.L. md RJ. Hoyle. 1972 Acute toxicity of yellow phmphom 14 Allantic ad 

(Wud mdua) and Atlantic salmon (Sdmo sdar) smolts. Jour. Fish. Rea Bd. of 

-29:l.aS. 

Hutchinmn, G.E 1957. A treatise on limnology. John Wiley and Sons, New York. 

Idler. D.R 1963. Coexistence of a fuhery and a major industy in Plaoentia Bay. 

Chembtry in Camda, 21(11):16. 

Imm B.G. IW. Toxicity of elementpry phosphorus. Jour. Water Poll. Convol Fed. 32: 

1312 

Jurgaard, P.Y. 1970. Ihe mle played by the Fiaheriea Research Board of Canada in the 

“d haring pbosphwus pollution crioia in Pkcentia Bay. Newfoundland. Circular No. 

1. FLheries Reseuch Bard, Atlantic Regional Mfioe, Hrli!u. Nova Smtk 

Yrkenthun, K.M. 197S. Toaud a derner quatic environment. US. Environmental 

Roteaion Agency, W.shington. D.C. 

Pe, D.L 1972 Obervations onmorulitiesof tenthicwgraismsiltercuntaminationof 

tbe bottom of &n!g Hybour, Aaoentia Bay, Newfoundl.od with ekmental phpb 

Rpea 181-186 in Eff- .of elemental phosphorus on mvine life. Fishries 

r!efsmb Board of csnadq cihr 2 

Volknweider. RA 1973. Input output nudela ScbwaebZ Hydml. 
SOLIDS (DISSOLVED) AND SALINITY 

REFERENCES CITED 

Apiculture Handbook h'o. 60.19%. Diagnosis and improvement of saline and alkali soils. 

L.A. Richards,ed. U.S.Government RintingOffice, Washingl0n.D.C. 

Bruvold, W.H., et al. 1969. Consumer -merit of mineral tante in domestic water. Jour. 

Amer. Water Workshn. 61:515. 

Capurm. LRA. 1970. Oceanopaphy for'practicing engineem. Barnes and Noble Inc., 

h'ea York. 

Griffith, W.H., Jr. 1963. Salt as a possible limiting factor to the Suisun Manh phearant 

population. Annual repon Delta Rsh and Wildlife Protection Study, Cooperative Study 

of California. 

hhman, J.A. 1964 Control of mmaion in water systems. Jour. Amer. Water Works 

hn. 56: 1009. 

Lodihart, EE, et al. 1955. The effect of water impurities on the flavor of brewed mffee. 

Fmd Research, 20: 5%. 

Moore, EW. 1952 Physiological effeets of the mnsumption of saline drinking water. 

National Res. Counal, Div. of Medid Seienw, Bull. San. Engr., and Environment. 

Appendix E. 

National Academy of Sciences, National Academy of Gngineering. 1974. Water quality 

criteria, 1972. US. Government Printing Office, Washington, D.C. 

National Research Council. 1954. Sodium resvicted diets. Publication 925, Food and 

Nutrition Board, Washington, D.C. 

National Technid Advisory Committee to the Secretary of the Interior. 1968. Water 

quality criteria. US. Government Rinting Office, Washian, D.C. 

Patterson, W.L., and RF. Banker. 1963. Effecta of higbly mineralized water on housebold 

plumbing and appliances. Jour. Am-. Water Worka hn. 60: 1060. 

Public Health Service. 1962 Drinking water atandarda, 1962 US. Government Printing 

Office, Waahian, D.C. 

Ra-n. D.S.. and J.E. Moore. lSi.4. The dine lakes of Saskatchewan. crnadian Jour. of 

Res.22: 141. 

Ricter. C.O., and A. MaeLean. 1939. Salt tante threshold of humans. Am. J. Phyiol. 126:l. 

Rouwfell, G.A., and W.H. Everhart 1953. Fisheryseience, its metbods and applications. 

'. 

John Wile?. and Sons. he., New York. 

Sawyer. C.N. 1960. Chembtrj for 3aniUry epgineem. Mffiraw-Hill Book Co., New York. 

Standard methods for the examination of waier and wastewater, 18thed.. 1971. Edited by 

Michael C. Rand. et a]. American Public Health hn., American Water Works Asan., 

Water Pollution Control Federation, Washington, D.C. 

SOLIDS (SUSPENDED, SETTLEABLE) AND TURBIDITY 

REFERENCES CITED 

Edberg. N. and Hofsten, B.V. 1973. Oxygen uptake of bottom sediment studied in-situ 

and in the Inboratory. Water Resemth, I: 1285. 

European. Inland fisheries Advisary Commiasion. 1965. Water quality criteria for 

European freshwater fiih, report on finely divided solids and inland fisheries. InL Jour. 

Air Water Poll. 9: 151. 

Gammon. J.R 1970. The effect of inorganic sediment on stmam biota Water Poll. Cont. 

Rea. Serieb, 18060 DWC WO. U.S. Environmental Proteetion Agency. US. Government Rinting Office. Washington, D.C. 

Mackenthun. K.M. 1973. Toward a cleaner aquatic envimnment US. Government 

Rinting Office, Washington. D.C. 

National Academy of Sciences, National Academy of Engineering. 1974. Water quality 

criteria, 1972 US. Government Printing Office, Washington, D.C. 

Standard methala for the eramination of water and wastewater, 1971.13th ed. Edited by 

Mictwl C. Rand, et al. American Public Health hn., American Water Works hn., 

Water Pollution Control Federation. 

Tmell. C.M., and A.R Gnufin. 1959. Some important hiologieal effecta of pollution 

often disregarded in StRam survey. Roceedings of the 0th RYdue industrial waste 

mnferrnce. Reprinted in Biology of water pollution, 1367. US. Depanmenr of 

Interior, Washingl0n. D.C. 

Tebo, L.B., Jr. 1955. Effects of siltation, resulting from improper logging, on the bottom 

fauna of a smll trout stream in the southern Appalachians. The RogRaaive fish 

Culturist. 11: €4. 
SULFIDE - HYDROGEN SULFIDE 

REFERENCES CITED 

Adelman. 1.R. and L.L. Smith. 1970. Effect of hy-3 sulfide on nwthern pike egga and 

Rmn. C.W.. and BJ. Follis. 1967. Effects of hydmren sulfide onchanneleaffuh Tram. 

aae fry. Trans. Amer. Fish. Soe 99: 501. 

, 

Amer Fiih. Scc 96 31 

dmon and mut Wash. Ikpt. Fish. Res. Bull.No. 5. 

Jom. 3 .R 1364. Fish and river pollution. Butteworth, London. 

National Wemy of Sciences. National Academy of Engineering. 1974. Water qUditY 

criteria 1972 U.S. Government Printing Offioe. WashingWn. D.C. 

bid. D.M., and L.L Smith, 1974a Chronic toxicity of hydrogen sulide to Gammnur 

pruddimnneus. Trans. Amer. Fish. Soe. Im. 

hid. D.M.. and L.L Smith. 1974b. Long termeffectsof hydmgensulfideonHc2ogen~ 

limbat0 (Ephemeroptera). Environmental Ecology. 

hid, D.M., and LL. Smith. 1974~. Facton influencing acute coxicity estimates of 

hydrogen sulfide to freshwater invertebrates Water Research 8. 

Smith, LL 1971. Influence of hydrogen sulfide on fish and arthmpods. Rojd 18050 

PCG, U.S. Environmental Rowtion Agency. WaahingWn, D.C. 

Smith, L.L.. and D.M. hid, 1972 Effects of hydrogen sulfide on fish eggs and fry. Water 

Reeesrch,G: 711. 

Smith, LL, Jr., and D.M. &id. 1974. Effwt of hy-n sulfide on development and 

sunrival of eight freshwater fish speck. Pagw 415-470 in J.H.S. Blaner, ed. Tbe early 

life history of fish. Sprinpr-Verlag. New York. 

Theede. H., et 81. 1969. Studies on the resistance of marine bottom invertebrates to 

oxygen deficiencies and hydrogen sulfrde. Mar. Biol. 2: 325. 

Van Horn, W.M. 1958. The effect of pulp and paper mill wastes on aquatic lik Roe. 

Ontario Indust. Wmte Cod. 5 60. 

TAINTING SUBSTANCES 

REFERENCES CITED 

Boyie. H.W. 1967. TaJte/dor mntamination of fish from theOhio River. Federal Water 

Pollution Conml Administration, Cincinnati. Ohio. 

National Academy of Sciences, National Academy of Engineering. 1974. Water quality 

criteria, 1972. U.S. Government Printing Office. Washington, D.C. 

Shumwsy, D.L.. and J.R. Palensky. 1973. Impairment of the flavor of fish by water 

pollutants. U.S. Environmental PmLection Agency, EPA-R.3-7M10, US. Government 

Rinting Office, Waahingwn, D.C. 

Thomas. N.A. 1973 Aasesament of fish flesh tinting substances. In J. Cairns and K.L. 

Dicksan, eds. Biological methods for the assessment of water quality. Amer. Soeiety for 

Testing and Materials. Tech. Publ. 528. Philadelphia 
TEMPERATURE 

REFERENCES CITED 

Ballentine. RK., and F.W. Kitmll. 1%. Obeervations of fed mlifom in several mnt 

stream pollution studies. RoeeedingJ of the Symposium on Fecal Coliform Bacterh in 

Water and Watewater, May 21-2,196s. Bureau of Sanituy Eqmeering, California 

State Department of Public Health. 

BlncL, E.C. 1953. Upper lethal temperatm of Mme British Columbia frsahwater fiihe. 

Jour. Fish. Res Bd. Can. 10 196. 

BretL J.R WI. Tempering venu acclimation in the planting of spgWedbvut Trura. 

Amer. Fish. Sx. 70: 391. 

BretL J.R 1%. Some principles in the thermal requirements of rubes. My Rev. 

Biol.31: 75. 

Brett J.R 1W. Thermal requirements of fwh-threo dsdes of study, 194&1970. In 

C.M. Tarzwell, ed. Biol*cal pmblem in water pollution. Public Hdth Senriee. U.S. 

Dept of Health. Education and Welfare. 

Burnson. B. 19S. Seamnal temperarm variations in relation to anter mtment. Jw. 

her. Water Warlo &an. 30: 793. 

Cairns. J.. Jr. 1456. Eifects of in& temperatures on qutk organkm Id4 

Warn. 1: 50. 

CalabRse. A 1968. Individual and mmbined effe of lulinity d temperature on 

edryoe and lmae of the mot clam, Mulink Mia (Spy). Biol. Bull. lS7,S: 417. 

Camp, T.R 1963. Water and iw impurities. Reinhold Publishing %., New Y& 

Chin, E. 1961. A trawl study of an estuarine nwsery wea in Gdveaton Bay ritb 

panicular referenoe la penaeid rhrimp. Ph.D. Dissenation, Univvsityof Wuhin%on 

Gmtloa, J.D.. Jr.. d C.G. BmWlout 1971. The effect of eydie templrtura on hd 

development in the mud aab, JZhitAmpowp~d haorririi. In D.I. Crisp, ed. Fd 

European marine biolcgy sympium Cambridge Univemity Rsa hdon 

Coutant. C.C. IW. Thermal pollution-biological effects: A review of the Literature of 

1967. Jour. Water Poll. Cont Fed. 40: 1M7. 

Coutant, C.C. 1968. Thermal pollution-biological effest.5: A review of the literature of 

1968. Jour. Water Poll. Cont Fed. 41: 1836. 

hunt, C.C. 1970, Thermal pllution-biological effects: A review of the Literature of 

1989. Jour. Water PoU. Cont Fed. 4: 1025. 

Couunt, C.C. 1971. Thermal pollution-biological effects: A review of the literature of 

1970. Jour. Water PoU. Cont Fed. 43: 1292 

Coumt, C.C. 1972 Biological ~peets of thermal pollution, 11. SdentifK hsis for rate 

temperature at~nduds nt power phnnta. CRC Critical Rev. in hviron Cant 8: L 

Coutant. C.C. 1975. Temperature selection hy &h-a fmr in pow- plant impact 

smeaumenfll In Symposium on the physical d biological effe on the environment 

of mling srstema d thermal discharp at nuclear pwer ations Intemstjorul 

Atomic hergy Agency. 

Coumt, C.C.. md C.P. coodyear. 1972 Water pollution-tkd pdlution: Areview of 

the literature of 1971 Jw. Water Poll. Cont Fed. 44: 1250. 

CoutlnL C.C., and H.A Pfuderer. 1978. Tbermal effe-litenture r&ew iaw Jour. 

Water Poll. Cont Fed. 45: EX. 

Couunt, C.C.. d HA Pfuderer. 1974. Tbernul effe-titmature review iaw Jour. 

Water Poll. Cont Fed 16: 1476. 

Crisp, DJ. 1957. Wfea of low temperature on the kdmg of myiDe dnvk Nature, 

179: 1x78. 

De Sylvr, D.P. 1983. l'hretkl mnsidemtions of tbe effd of heated effluents on 

mvine f'sbes In PA KreJel d F.C. Mu, eds. BioloBiaJ .spgu of thermal 

pollution V.aderbilt Uuivemity Freat, NMhville. Tern 

dtvluuiug, V.L 1971. Tbe effecb of food deprivation d vlinity ehnges on repduetive fm&m in the rstuuine pbiid fmh, C;Uiehuy# m'mbilia. Biol. Bull 141: 4.58. 

devluning, V.L l977. The effe of temperature ud pbotoperiod on reproductive 

cycling of tbe mtuuine pbiid fMh, Gilliehuya nirrdilu. fLhry Bull. 70,4: 1151. 

Dow. RL W73. Fluctutioua in mvine apgies .buwhce during clinutic cyck Mu, 

Teh Sx. Jour. 7k 5%. 

Environmental Resevch Lboramy. 1976. prmdwra for developing tempeatwe 

miteria for frahwp.ter f& Emlogid Reaeych series. repfi in -tion ERL' 

Duluth. Mjm 

PLdenl Wrtpr Pollution Contml Administnrion l967. Tempenture ud qutic life. 

L.bontmy laveatigatiorm No. 6. Technical Ad.risory lad Invmtigatiom Branch, 

Cidti. Ohio. 

Fry, F.EJ. 1967. Rmpomz of vertebrate poiLilotbenns to tewersture In AH. Rose. ed. 

--h=xa eiwr Vorbehandlung nit koast.nten urd wghaelnden 

tempentauen uti die hiWemistenr von Gommonu sdiw uod I&ta kJuim &. 

Biol. 15: 12 

Clude, J.B. ISM The effd of temperature d psd.tora on the .bundance of the 

mfubell drm,dlyoamk,in New Fh&d. TMb Amer. Fish Soc & 19. 

I 

' Rcm,NewYok 
Glynn. P.W. 1963. b modities of eehiwids andocberrpefllrtog.nismsmi~nt 

6th midday. low water erposursa in Puerto Rieo. Mar. Biol. 3 226. 

Gouda. J.G. 1~iZ hM[ vanition in tbe reapow of &ne popuhtioy to kted 

warn in tbe vidnity of m gewratang phnt Ph.D. DLssenrtaon, Umversty of 

RbodekLnd. 

Hd. SA et J. 1967. Contml techniques for mrgulrtiobfdbation Jwr. Amer. Water 
IT 

I .I' 

I 

E! ! ! 

, ... --- ... ~ 

.., 

J~VOL 

.,I.. 

7., 
I I 

! 
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY 

OFFICE OF WATER REGULATIONS AND STANDARDS 

CRITERIA AND STANDARDS DIVISION 

UPDATE t2 

to 

"QUALITY CRITERIA FOR WATER 1986" 

May 1, 1987 

This is the second update to the EPA document "Quality Criteria 

for Water 1986". Included in this package are criteria summaries 

for contaminants that were recently revised as well as a criteria 

summary for a new contaminant. Several hand corrections are also 

included. 

@ Revised New Hand corrections 

INDEX CHLORPYRIFOS AMMONIA 

SUMMARY CHART CYANJDE 

NICKEL CHLORINATED ETHANES 

P ARATH I ON 

PENTACHLOROPHENOL 

TOXAPHENE 

ZINC 

Directions: 

a. Replace sections that have been revised with 

b. Insert new section alphabetically. 

c. Make the identified hand corrections. 

new sections. 

For additional information contact: 

EPA's Criteria and Standards Division 

at (202) 475-7315. 
I HAND CORRECTIONS .. 

.. .. 

oaia - page 5, center page, third line of equation should be 

changed from 

.' . "FPH = 1 : 8 < pH < 9" 

to . ., 

; 8 5 pH < 9" "FPH = 1 - ' '. .. 

................................................................. .. 

Cyanide - last page, include the following after the last line 

.. .. "(45 F-R- 79318,NOV. 28, 1980) (50 F.R. 30784, July 29. 1985)" 

."SEE APPENDIX A FOR METHODOLOGY" 

., .. ...,. 

-. ':.'Chlorinated Ethanes - last page, change 

.. 

1 

"1.03 ug/l" 

to 

"1.03 g/L" 

, 
INDEX 

INTRODUCTION 

SUMMARY CHART 

NTIS No. 

PB 81-117269 

PB 81-117277 

PB 81-117285 

PB-263943 

PB-263943 

PB 81-117301 

PB 85-227114 

PB 81-117319 

PB 85-227445 

PB 81-117335 

PB-263943 

PB 81-117293 

PB 81-117343 

PB 81-117350 

PB-263943 

PB 85-227031 

PB 81-117376 

PB 81-1’17384 

PB 81-117392 

PB 81-117400 

PB 81-117426 

PB 85-227429 

PB 81-117434 

PB 81-117418 

PB 81-117442 

PB-263943 

PB 87-105359 

PB 85-227478 

PB 81-117459 

PB-263943 

PB 85-227023 

PB 85-227460 

PB 81-117491 

PB-263943 

PB 81-117509 

PB 81-117517 

PB 81-117525 

PB 81-117533 

PB 86-158-045 €i PB-263943 

Acenaphthene 

Acrolein 

Acrylonitrile 

Aesthetics 

Alkalinity 

AldrinDieldrin 

Ammonia 

Antimony 

Arsenic 

Asbestos 

Bacteria 

Barium 

Benzene 

Benzidine 

Beryllium 

Boron 

Cadmium 

Carbon Tetrachloride 

Chlordane 

Chlorinated Benzenes 

Chlorinated Ethanes 

Chlorinated Naphthalenes 

Chlorine 

Chlorinated Phenols 

Chloroalkyl Ethers 

Chloroform 

Chlorophenoxy Herbicides 

Chlorpyrifos 

Chromium 

2-Chlorophenol 

Color 

Copper 

Cyanide 

DDT and Metabolites 

Demeton 

Dichlorobenzenes 

Dichlorobenzidine 

Dichloroethylenes 

2.4. - DichloroDhenol 

Dichloropropanes/Dichloropropenes 

2.4. - Dimethylphenol 

Dinitrotoluene 

Diphenylhydrazine 

Endosulfan 

Endrin 

Ethylbenzene 

Fluoranthene 

Gasses, Total Dissolved 

PB 81-117541 

PB 81-117558 

PB 81-117566 

PB 81-117731 

PB 81-117574 

PB 81-117582 

PB 81-117590 

PB-263943 

PB 8i-ii76oa 
Guthion 

Haloethers 

Halomethanes 

Hardness 

Heptachlor 

Hexachlorobutadiene 

Hexachlorocyclohexane 

Hexachlorocyclopentadiene 

Iron 

Isophorone 

Lead 

Ma la thion 

Manganese 

Mercury 

Methoxychlor 

Mirex 

Naphthalene 

Nickel 

Nitrates. Nitrites 

Nitrobenzene 

Nitrophenols 

Nitrosamines 

Oil and Grease 

Oxygen, Dissolved 

Parathion 

Pentachlorophenol 

Ph 

Phenol 

Phosphorus 

Phthalate Esters 

Polychlorinated Biphenyls 

Polynuclear Aromatic Hydrocarbons 

Selenium 

Silver 

Solids (Dissolved) h Salinity 

Solids (Suspended) h Turbidity 

Sulfides, Hydrogen Sulfide 

Tainting Substances 

Temperature 

2.3.7.8-Tetrachlorodibenzo-p-dioxin 

Tetrachloroethylene 

Thallium 

Toluene 

Toxaphene 

Trichlorocthylene 

Vinyl Chloride 

Zinc 

PB-263943 

PB 81-117616 

PB 81-117624 

PB-263943 

PB 81-117632 

PB 81-117640 

PB 81-117657 

PB 81-117665 

PB-263943 

PB 81-117673 

PB 85-227437 

PB-263943 

PB-263943 

PB 85-227452 

PB-263943 

PB-263943 

PB 81-117707 

PB 87-105359 

PB-263943 

PB 81-117723 

PB 81-117749 

PB 81-117756 

PB-263943 

PB 86-208253 

PB 87-105383 

PB 87-105391 

PB-263943 

PB 81-117772 

PB-263943 

PB 81-117780 

PB 81-117798 

PB 81-117806 

PB 81-117814 

PB 81-117822 

PB-263943 1 

PB-263943 

PB-263943 

PB-263943 

PB-263943 

EPA # 440/5-84-007 

PB 81-117830 

PB 81-117848 

PB 81-117855 

PB 87-105375 

PB 81-117871 

PB 81-117889 

PB 87-153581 

APPENDIX A Methodology for Developing Criteria 

APPENDIX B Methodology for Developing Criteria 

APPENDIX C Methodology €or Developing Criteria 

BIBLIOGRAPHY a 
CHLORPYRIFOS 

Summary 

The acute values for eighteen freshwater species in fifteen 

genera range from 0.11 ug/L for an amphipod to greater than 806 

ug/L for two fishes and a snail. The bluegill is the most 

acutely sensitive fish species with an acute value of 10 ug/L, 

but seven intervetebrate genera are more sensitive. Smaller 

organisms seem to be more acutely sensitive than larger ones. 

Chronic toxicity data are available for one freshwater 

species, the fathead minnow. Unacceptable effects occurred in 

second generation larvae at 0.12 ug/L, which was the lowest 

concentration tested. The resulting acute-chronic ratio was 

greater than 1,417. 

Little information is available on the toxicity of 

chlorpyrifos to freshwater plants, although algal blooms 

frequently follow field applications of chlorpyrifos. The only 

available bioconcentration test on chlorpyrifos with a freshwater 

species was with the fathead minnow and resulted in a bioconcentration factor of 1,673. 

The acute toxicity of chlorpyrifos has been determined for 15 

species of saltwater animals in 12 genera with the acute values 

ranging from 0.01 ug/L for the Korean shrimp, Palaemon 

macrodactvlus, to 1.911 ug/L for larvae of the eastern oyster, 

Crassostrea virginica. Arthropods are particularly sensitive to 

chlorpyrifos. Among the 10 species of fish tested, the 96-hr 

LC50s range from 0.58 ug/L for striped bass to 520 ug/L for gulf 
toadfish. Fish larvae are more sensitive than other life stages. 

Growth of the mysid, Mysidolssis bahia, was reduced at 0.004 ug/L 

in a life-cycle test. In early life-stage tests, the California 

grunion, Leuresthes tenuis, was the most sensitive of the six 

a. 

fishes, with growth being reduced at 0.30 ug/L. Of the seven 

acute-chronic ratios that have been determined with saltwater 

species, the five lowest range from 1.374 to 12.50, whereas the 

highest is 228.5. 

Concentrations of chlorpyrifos affecting six species of 

saltwater phytoplankton range from 138 to 10,000 ug/L. BCFs 

ranged from 100 to 5,100 when the gulf toadfish was exposed to 

.oncentrations increasing from 1.4 to 150 ug/L. Steady-state 

BCFs averaged from 100 to 757 for five fishes exposed in early 

life-stage tests. 

National Criteria 

The procedures described in the "Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses" indicate, that except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of 

chlorpyrifos does not exceed 0.041 ug/L more than once every 

three years on the average and if the one-hour average 

concentration does not exceed 0.083 ug/L more than once every 

three years on the average. 

The procedures described in the "Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 
Aquatic Organisms and Their Uses" indicate, that except possibly 

a where a locally important species is very sensitive, saltwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of 

chlorpyrifos does not exceed 0.0056 ug/L more than once every 

three years on the average and if the one-hour average 

concentration does not exceed 0.011 ug/L more than once every 

three years on the average. 

Three yeaks is the Agency's best scientific judgment of the 

average amount of time aquatic ecosystems should be provided. 

between excursions. The resiliences of ecosystems and their 

abilities to recover differ greatly, however, and site-specific 

allowed excursion frequencies may be established if adequate 

justification is provided. 

a Use of criteria for developing water quality-based permit 

limits and for designing waste treatment facilities requires 

selection of an appropriate wasteload allocation model. Dynamic 

models are preferred for the application of these criteria. 

Limited data or other consideretions might make their use 

impractical, in which case one must rely on a steady-state model. 

(51 F.R. 43665, December 3, 1986) 

SEE APPENDIX A FOR METHODOLOGY 
a 

Summary 

Acute values with twency-one freshwater species in 18 genera 

range from 1,101 ug/L for a cladoceran to 43,240 ug/L for a fish. 

Fishes and invertebrates are both spread throughout the range of 

sensitivity. Acute values with four species are significantly 

correlated with hardness. Data are available concerning the 

chronic toxicity of nickel to two invertebrates and two fishes in 

freshwater. Data available for two species indicate that chronic 

toxicity decreases as hardness increases. The measured chronic 

values ranged from 14.77 ug/L with DaDhnia maqna in soft water to 

526.7 ug/L with the fathead minnow in hard water. Five acutechronic ratios are available for two species in soft and hard 

water and range from 14 to 122. a Nickel appears to be quite toxic to freshwater algae, with 

concentrations as low as 50 ug/L producing significant 

inhibition. Bioconcentration factors for nickel range from 0.8 

for fish muscle to 193 for a cladoceran. I 

Acute values for twenty-three saltwater species in twenty 

genera range from 151.7 ug/L with juveniles of a mysid of to 

1,100,000 ug/L with juveniles and adults of a clam. The acute 

values for the four species of fish range from 7,598 to 350.000 

ug/L. The acute toxicity of nickel appears to be related to 

salinity, but the form of the relationship appears to be speciesdependent. 

Mysidopsis bahia is the only saltwater species with which an 

acceptable chronic test has been conducted on nickel. Chronic 0 
exposure to 141 ug/L and greater resulted in reduced survival and 

reproduction. The measured acute-chronic ratio was 5.478. 01 

Bioconcentration factors in saltwater range from 261.8 with a 

oyster to 675 with a brown alga. 

National Criteria 

The procedures described in the "Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses" indicate, that except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of nickel (in 

(0.8460[ In ug/L) does not exceed the numerical value given by e 

(hardness)1+1.1645) more than once every three. years on the 

average and if the one-hour average concentration (in ug/L) does 

not exceed the numerical value given by e (0.846011n 0) 

more than once every three years on the (hardness)1+3.3612) 

average. For example, at hardnesses of 50. 100, and 200 mg/L as 

CaC03 the four-day average concentrations of nickel are 88, 160. 

and 280 ug/L, respectively, and the one-hour average 

concentrations are 790, 1400. and 2500 ug/L. 

The procedures described in the "Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses" indicate, that except possibly 

where a locally important species is very sensitive, saltwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of nickel does 

not exceed 8.3 ug/L more than once every three years on the 

average and if the one-hour average concentration does not exceed 0) 
a 

75 ug/L more than once every three years on the average. 

"Acid-soluble'' is probably the best measurement at present 

for expressing criteria for metals and the criteria for nickel 

were developed on this basis. However, at this time, no EPA 

approved method for such a measurement is available to implement 

criteria for metals through the regulatory programs of the Agency 

and the States. The Agency is considering development and 

approval of a method for a measurement such as "acid-soluble." 

Until one is approved, however, EPA recommends applying criteria 

for metals using the total recoverable method. This has two 

impacts: (1) certain species of some metals cannot be measured 

because the total recoverable method cannot distinguish between 

individual oxidation States, and (2) in some cases these criteria 

might be overly protective when based on the total recoverable 

method . 

Three years is the Agency's best scientific judgment of the 

average amount of time aquatic ecosystems should be provided 

between excursions. The resiliences of ecosystems and their 

abilities to recover differ greatly, however, and site-specific 

allowed excursion frequencies may be established if adequate 

justification is provided. 

Use of criteria for developing water quality-based permit 

limits and for designing waste treatment facilities requires 

selection of an appropriate wasteload allocation model. Dynamic 

models are preferred for the application of these criteria. 

Limited data or other considerations might make their use 

impractical, in which case one must rely on a steady-state model. 
(51 F.R. 43665, December 3, 1986) 

SEE APPENDIX A FOR METHODOLOGY 
PARATHION 

Summary 

The acute values for thirty-seven freshwater species in 

thirty-one genera range from 0.04 ug/L for an early instar of a 

crayfish, Orconectes nais, to 5,230 ug/L for two species of 

tubificid worms. For Daphnia mama, the chronic value and acutechronic ratio are 0.0990 ug/L and 10.10 respectively. Chronic 

toxicity values are available for two freshwater fish species, 

the bluegill and the fathead minnow, with chronic values of 0.24 

ug/L and 6.3 ug/L, and acute-chronic ratios of 2.121 and 79.45. 

respectively. Two freshwater algae were affected by toxaphene 

concentrations of 30 and 390 ug/L, respectively. 

Bioconcentration factors determined with three fish species 

ranged from 27 to 573. 

The acute values that are available for saltwater species are 

11.5 and 17.8 ug/L for the Korean shrimp, Palaemon macrodactvlus. 

and 17.8 ug/L for the striped bass, Morone saxatilis. No data 

are available concerning the chronic toxicity of parathion to 

saltwater species, toxicity to saltwater plants, or 

bioaccumulation by saltwater species. Some data indicate that 

parathion is acutely lethal to commercially important saltwater 

shrimp at concentrations as low as 0.24 ug/L. Measurement of 

acetylcholinesterase (AChE) in fish tissue might be useful for 

diagnosing fish kills caused by parathion. 

National Criteria 

The procedures described in the “Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 
Aquatic Organisms and Their Uses“ indicate, that except possibly 

where a locally important species is very sensitive, freshwater a 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of parathion 

does not exceed 0.013 ug/L more than once every three years on 

the average and if the one-hour average concentration does not 

exceed 0.065 ug/L more than once every three years on the 

average. 

The procedures described in the “Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses” require the availability of 

specified data for the derivation of a criterion. A saltwater 

criterion for parathion cannot be derived because very few of the 

required data are available. 

Three years is the Agency’s best scientific judgment of the .I 

average amount of time aquatic ecosystems should be provided 

between excursions. The resiliences of ecosystems and ‘their 

abilities to recover differ greatly, however, and site-specific 

allowed excursion frequencies may be established if adequate 

justification is provided. 

Use of criteria for developing water quality-based permit 

limits and for designing waste treatment facilities requires 

selection of an appropriate wasteload allocation model. Dynamic 

models are preferred for the application of these criteria. 

Limited data or other considerations might make their use 

impractical, in which case one must rely on a steady-state model. 

(51 F.R. 43665, December 3, 1986) 

SEE APPENDIX A FOR METHODOLOGY 
PENTACHLOROPHENOL (PCP) 

Summary 

The acute and chronic toxicity of PCP to freshwater animals 

increased as pH and dissolved oxygen concentration of the water 

decreased. Generally, toxicity also increased with increased 

temperature. The estimated acute sensitivities of 36 species at 

pH = 6.5 ranged from 4.355 ug/L for larval common carp to >43,920 

ug/L for a crayfish. At pH = 6.5, the lowest and highest 

estimated chronic values of 11.835 and 79.06 ug/L, respectively, 

were obtained with different cladoceran species. Chronic 

toxicity to fish was affected by the presence of impurities, with 

industrial-grade PCP being more toxic than purified samples. 

Mean acute-chronic ratios for six freshwater species ranged from 

0.8945 to >15.79, but the mean ratios for the four most acutely 

sensitive species only range 0.8945 to 5.035. Freshwater algae 

were affected by concentrations as low as 7.5 ug/L, whereas 

vascular plants were affected at 189 ug/L and above. 

Bioconcentration factors ranqed from 7.3 to 1,066 for three 

species of fish. 

Acute toxicity values from tests with 18 species of saltwater 

animals, representing 17 genera, range from 22:63 ug/L for late 

yolk-sac larvae of the Pacific herring, Clupea narenqus pallasi, 

to 18,000 ug/L for adult blue mussels, Mvtllus edulis. The 

embryo and larval stages of invertebrates and the late larval 

premetamorphosis stage of ash appear to be the most sensitive 

0 life stages to PCF. With few exceptions, Msh are more sensitive 
than 

slight 

invertebrates 

effect on the toxicity 

to PCP. Salinity, temperature, and 

of PCP to some saltwater 

pR 

animals. 

have a .

Life-cycle toxicity tests have been conducted with the 

sheepshead minnow, Cmrinodon variegatus. and the polychaete 

worm, Ophrvotrocha diadema. The chronic value for the minnow is 

64.31 ug/L and the acute-chronic ratio is 6.873. Unfortunately, 

no statistical analysis of the data from the test with the worm 

is available. 

The EC506 for saltwater plants range from 17.40 ug/L for the 

diatom, Skeletonema costatum, to 3.600 ug/L for the green alga, 

Dunaliella tertiolecta. Apparent steady-state BCFs are available 

for the eastern oyster. Crassostrea virginica, and two saltwater 

fishes and range from 10 to 82. 

National Criteria 

The procedures described in Fe "Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses" indicate, that except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration (in ug/L) of 

pentachlorophenol does not exceed the numerical value given by 

e 11.005(pH)-5-2901 more than once every three years on the 

average and if the one-hour average concentration (in ug/L) does 

not exceed the numerical value given by e [l.OOS(pH)-4.8301 

than once every three years on the average. For example, at pH = 

6.5. 7.8, and 9.0 the four-day average concentrations 

pentachlorophenol are 3.5, 13, and 43 ug/L, respectively, and the 
one-hour average concentrations are 5.5, 20, and 68 ug/L. At pH 

= 6.8, a pentachlorophenol concentration of 1.74 ug/L caused a 

50% reduction in the growth of yearling sockeye salmon in a 56- 

day test. 

a 

The procedures described in the “Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses“ indicate, that except possibly 

where a locally important species is very sensitive, saltwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of 

pentachlorophenol does not exceed 7.9 ug/L more than once every 

three years on the average and if the one-hour average 

concentration does not exceed 13 ug/L more than once every three 

years on the average. 

Three years is the Agency’s best scientific judgment of the 

average amount of time aquatic ecosystems should be provided 

between excursions. The resiliences of ecosystems and their 

abilities to recover differ greatly, however, and site-specific 

allosred excursion frequencies may be established if adequate 

justification is provided. 

Use of criteria for developing water quality-based permit 

limits and for designing waste treatment facilities requires 

selection of an appropriate wasteload allocation model. Dynamic 

models are preferred for the application of these criteria. 

Limited data or other considerations might make their we 

impractical, in which case one must rely on a steady-state model. 

(51 F.R. 43665, December 3, 1986) 

SEE APPENDIX A FOR METHODOLOGY 
TOXAPHENE 

Summary 

The acute sensitivities of 36 freshwater species in 28 genera 

ranged from 0.8 ug/L to 500 ug/L. Such important fish species as 

the channel catfish, largemouth bass, chinook and coho salmon, 

brook, brown and rainbow trout, striped bass, and bluegill had 

acute senitivities ranging from 0.8 ug/L to 10.8 ug/L. Chronic 

values for four freshwater species range from less than 0.039 

ug/L for the brook trout to 0.1964 ug/L for the channel catfish. 

The growth of algae was affected at 100 to 1,000 ug/L, and 

bioconcentration factors from laboratory tests ranged from 3.100 

to 90,000. Concentrations in lake trout in the Great Lakes have 

frequently exceeded the U.S. FDAJaction level of 5 mg/kg, even 

though the concentrations in the water seem to be only 1 to 4 

ng/L. These concentrations in lake water are thought to have 

resulted from toxaphene being transported to the Great Lakes from 

remote sites, the locations of which are not well known. 

a 

The acute toxicity of toxaphene to 15 species of saltwater 

animals ranges from 0.53 for pinfish, Laqodon rhomoides. to 

460.000 ug/L for the adults of the clam, Ranqia cuneata. Except 

for resistant species tested at concentrations greater than 

toxaphene's water solubility, acute values for most species were 

within a factor of 10. The toxicity of toxaphene was found to 

decrease slightly with increasing salinity for adult blue crabs. 

Callinectes sapidus, whereas no relationship between toxicity and 

salinity was observed with the three spine stickleback, 
Gasterosteus aculeatus. Temperature significantly affected the 

toxicity of toxaphene to blue crabs. 

Early life-stage toxicity tests have been conducted with the 

sheepshead minnow, Cmrinodon varieqatus, and the longnose 

killifish, Fundulus similis, whereas life-cycle tests have been 

conducted with the sheepshead minnow and a mysid. For the 

sheepshead minnow, chronic values of 1.658 ug/L from the early 

life-stage test and 0.7141 ug/L from the life-cycle toxicity test 

are similar to the 96-hr LC50 of 1.1 ug/L. Killifish are more 

chronically sensitive with effects noted at 0.3 ug/L. In the 

life-cycle test with the mysid, no adverse effects were observed 

at the highest concentration tested, which was only slightly 

below the 96-hr LC50, resulting in an acute-chronic ratio of 

1.132. 

Toxaphene is bioconcentrated by an oyster, Crassostrea 01 

virsinica, and two fishes, C. variegatus and z- similis, to 

concentrations that range from 9,380 to 70.140 times that in the 

test solution. 

National Criteria 

The procedures described in the "Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses" indicate, that except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of toxaphene 

does not exceed 0.0002 ug/L more than once every three years on 

the average and if the one-hour average concentration does not 
exceed 0.73 ug/L more than once every three years on the average. 

0 If the concentration of toxaphene does exceed 0.0002 ug/L. the 

edible portions of consumed species should be analyzed to 

determine whether the concentration of toxaphene exceeds the FDA 

action level of 5 mg/kg. If the channel catfish is as acutely 

sensitive as some data indicate it might be, it will not be 

protected by this criterion. 

The procedures described in the “Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses“ indicate, that except possibly 

where a locally important species is very sensitive, saltwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of toxaphene 

does not exceed 0.0002 ug/L more than once every three years on 

the average and if the one-hour average concentration does not 

exceed 0.21 ug/L more than once every three years on the average. 

If the concentration of toxaphene does exceed 0.0002 ug/L, the 

edible portions of consumed species should be analyzed to 

determine whether the concentration of toxaphene exceeds the FDA 

action level of 5 mg/kg. 

e 

I 

Three years is the Agency’s best scientific judgment of the 

average amount of time aquatic ecosystems should be provided 

between excursions. The resiliences of ecosystems and their 

abilities to recover differ greatly, however, and site-specific 

allowed excursion frequencies may be established if adequate 

justification is provided. 

Use of criteria for developing water quality-based permit 

0 limits and for designing waste treatment facilities requires 
selection of an appropriate wasteload allocation model. Dynamic 

models are preferred for the application of these criteria. 

Limited data or other considerations might make their use 

impractical, in which case one must rely on a steady-state model. 

(51 F.R. 43665, December 3, 1986) 

SEE APPENDIX A FOR METHODOLOGY 

0) 
Summary 

Acute toxicity values are available for 43 species of 

fresl-rater animals and data for eight species indicate that acute 

toxicity decreases as hardness increases. When adjusted to a 

hardness of 50 mg/L, sensitivities range from 50.70 ug/L for 

CeriodaDhnia reticulata to 88,960 ug/L for a damselfly. 

Additional data indicate that toxicity increases as temperGture 

increases. Chronic toxicity data are available for nine 

freshwater species. Chronic values for two invertebrates ranged 

from 46.73 ug/L for DaDhnia mama to >5,243 ug/L for the 

caddisfly, Clistoronla magziifica. Chronic values for seven fish 

species ranged from 36.41 ug/L for the flagfish, Jordanella 

9 floridae, to 854.7 ug/L for the brook trout, Salvelinus 

fontinalis. Acute-chronic ratios ranged from 0.2614 to 41.20. 

but the ratios for the sensitive species were all less than 7.3. 

The sensitivity range of freshwater plants to zinc is greater 

than that for animals. Growth of the alga, Selenastrum 

capriocornutum, was inhibited by 30 ug/L. On the other hand, 

with several other species of green algae, 4-day EC5Os exceeded 

200,000 ug/L. Zinc was found to bioaccumulate in freshwater 

animal tissues from 51 to 1,130 times the concentration present 

in the water. 

Acceptable acute toxicity values for zinc are available for 

33 species of saltwater animals including 26 invertebrates and 7 

fish. LCSOs range from 191.5 ug/L for cabezon, ScorDanichthvs 

a 
marmoratus to 320.400 ug/L for adults of another clam, Macoma 

balthica. Early life stages of saltwater invertebrates and 01 

fishes are generally more sensitive to zinc than juveniles and 

adults. Temperature has variable and inconsistent effects on the 

sensitivity of saltwater invertebrates to zinc. The sensitivity 

of saltwater vertebrate animals to zinc deyreases with increasing 

salinity, but the magnitudG of the effect is species-specific. 

A life-cycle test with the mysid, Mvsidopsia bahia, found 

unacceptable effects at 120 ug/L, but not at 231 ug/L, and the 

acute-chronic ratio was 2.997. Seven species of saltwater plants 

were affected at concentrations ranging from 19 to 10,100 ug/L. 

Bioaccumulation data for zinc are available for seven species of 

saltwater algae and five species of saltwater animals. Steadystate zinc bioconcentration factors for the twelve species range 

from 3.692 to 23.820. 

National Criteria 

The procedures described in the "Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses" indicate, that except possibly 

where a locally important species is very sensitive, freshwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of zinc (in 

(0.847311n ug/L) does not exceed the numerical value given by e 

more than once every three years on the (hardness)l+0.7614) 

average and if the one-hour average concentration (in ug/L) does 

(0.8473 [ In not exceed the numerical value given by e 

more than once every three years on the (hardness)]+0.8604) 
0 

average. For example, at hardnesses of 50, 100, and 200 mg/L as 

CaC03. the four-day average concentrations of zinc are 59, 110 and 

190 ug/L, respectively, and the one-hour average concentrations 

are 65, 120, and 210 ug/L. If the striped bass is as sensitive 

as some data indicate, it will not be protected by this 

criterion. 

The procedures described in the "Guidelines for Deriving 

Numerical National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses " indicate, that except possibly 

where a locally important species is very sensitive, saltwater 

aquatic organisms and their uses should not be affected 

unacceptably if the four-day average concentration of zinc does 

not exceed 86 ug/L more than once.every three years on the 

average and if the one-hour average concentration does not exceed 

a 95 ug/L more than once every three years on the average. 

"Acid-soluble'' is probably the best measurement at present 

for expressing criteria for metals and the criteria for zinc were 

developed on this basis. However, at this time no FPA approved 

method for such a measurement is available to implement criteria 

for metals through the regulatory programs of the Agency and the 

States. The Agency is considering development and approval of a 

method for a measurement such as "acid-soluble.'' Until one is 

approved, however, EPA recommends applying criteria for metals 

using the total recoverable method. This has two impacts: (1) 

certain species of Some metals cannot be measured because the 

total recoverable method cannot distinguish between individual 

oxidation States, and (2) in some cases these criteria might be 

overly protective when based on the total recoverable method. 
Three years is the Agency’s best scientific judgment of the 

average amount of time aquatic ecosystems should be provided a 

between excursions. The resiliences of ecosystems and their 

abilities to recover differ greatly, however, and site-specific 

allowed excursion frequencies may be established if adequate 

justification is provided. 

Use of criteria for developing water quality-based permit 

limits and for designing waste treatment facilities requires 

selection of an appropriate wasteload allocation model. Dynamic 

models are preferred for the application of these criteria. 

Limited data or other considerations might make their use 

impractical, in which case one must rely on a steady-state model. 

(52 F.R. 6213, March 2. 1987) 

SEE APPENDIX A FOR METHODOLOGY