Sinn
Ertvironrninui Protection
Agtncy
Offict of Wattr
Btgul»toni me Stiooircti
WMftington. DC 20X60
X-/EPA
May 1, 19B6
EPA 440/5-86-001
QUALITY CRITERIA
for
WATER
1986
-------
BIBLIOGRAPHIC INFORMATION
PB87-226759
Quality Criteria for Water, 1986.
1 May 86
PERFORMER: Environmental Protection Agency, Washington, DC.
Office of Water Regulations and Standards.
EPA/440/5-86/001
Also available from Supt. of Docs.
Section 304(a) (1) of the Clean Water Act (33 U.S.C. 1314(a)
(1} requires the Environmental Protection Agency (EPA) to
publish and periodically update ambient water quality
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. 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
KEYWORDS: *Water quality, *Criteria, *Water pollution,
*Ambient water .quality criteria, *Quality
criteria.
Available from the National Technical Information Service,
SPRINGFIELD, VA. 22151
PRICE CODE: PC A18/MF A01
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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
publish and periodically update ambient water quality 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 *al«> by the Superintendent of DocMments, U,S Government Printing OfHc
oR, &C. 20402
-------
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 for pollutants listed as
toxic under section 307(a)(i) 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 Qual^^it^ Criteria for
Water 1986 was assembled. This document includes summaries of
all the contaminants for which EPA has developed criteria recom-
mendations (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
-------
-------
3
criteria documents containing all the data used to develop the
criteria recommendations summarized herein and the "Red Book",
including complete bibliographies are available only from:
National Technical Information Service
5285 Port Royal load
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 CrJ^ter^ia fgr, E§.te£i iEJL§ is designed to be easily
updated to reflect EPA's continuing work to present the latest
scientific information and practices. Our planned schedule for
future criteria development in the next few years is attached for
your information.
The Agency is currently 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:
-------
Chemical
1. Cyanide
2. Ethylbenzene
3, Nitrobenzene
4. Phenol
5. Toluene
1980 WQC
200 ug/L
1.4 mg/L
li.8 mg/L
3.S mg/L
14.3 ng/L
Draft RfD
.02 ing/kg/day
.01 ing/kg/day
.0005 »g/kg/day
0.l ing/kg/day
0.3 mg/kg/day
FOR FURTHER 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 K. conlon
f I Acting Director,
' Office of Water Regulations
and Standards
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DRAFT CRITERIA DOCUMENTS TO BE PROPOSED
LATE FY 86/EARLY FY 87
Diethyhexylphthalate
1,2,4, Trichlorobenzene
Silver
Phenanthrene
2,4,5, Trichlorophenol
Organotins
Tributyltin
Selenium (no saltwater criteria)
Hexachlorobenzene
Antimony III
Acrolein (no saltwater criteria)
LATE FY 87/EARLY 8£
Thallium (no saltwater criteria)
Tetrachloroethylene (no saltwater criteria)
Phenol
Toluene
Chloroform (no saltwater criteria)
Analine
Acrylontrile
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 PB 86
Barium
Benzene
Benzidine
Beryllium
Boron
Cadmium
Carbon Tetrachloride
Chlordane
Chlorinated Benzenes
Chlorinated Ethanes
Chlorinated Naphthalenes
Chlorine
Chlorinated Phenols
Chloroalkyl Ethers
Chloroform
Chlorophenoxy Herbicides
Chromium
2-Chlorophenol
Color
Copper
Cyanide
DDT and Metabolites
Demeton
Dichlorobenzenes
Dichlorobenzidine
Dichloroethylenes
2,4, - Dichlorophenol
Dichloropropanes/Dichloropropenes
2,4, - Dimethylphenol
Dinitrotoluene
DiphenyIhydraz ine
Endosulfan
Endrin
Ethylbenzene
riuoranthene
Gasses, Total Dissolved
Guthion
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
-158-045 & PB-263943
PB-263943
PB 81-117293
PB 81-117343
PB 81-117350
PB-263943
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-263943
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 81-117541
PB 81-117558
PB 81-117566
PB 81-117731
PB 81-117574
PB 81-117582
PB 81-117590
PB 81-117608
PB-263943
PB-263943
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Haloethers
Halomethanes
Hardness
Heptachlor
Hexachlorobutadiene
Hexachlorocyclohexane
Hexachlorocyclopentadiene
Iron
Isophorone
Lead'
Malathion
Manganese
Mercury
Methoxychlor
Mi rex
Naphthalene
Nickel
Nitrates, Nitrites
Nitrobenzene
Nitrophenols
Nitrosamines
Oil and Grease
Oxygen, Di ssolved
Parathion
Pentachlorophenol
Ph
Phenol
Phosphorus
Phthalate Esters
Polychlorinated Biphenyls
Polynuclear Aromatic Hydrocarbons
Selenium
Silver
Solids (Dissolved) & Salinity
Solids (Suspended) £ Turbidity
Sulfides, Hydrogen Sulfide
Tainting Substances
Temperature
2,3,7,8-Tetrachlorodiben20-p-dioxin
Tetrachloroethylene
Thallium
Toluene
Toxaphene
Trichloroethylene
Vinyl Chloride
Zinc
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 81-117715
PB-2639*43
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
PB-263943
EPA # 440/5-84-007
PB 81-117830
PB 81-117848
PB 81-117855
PB 81-117863
PB 81-117871
PB 81-117889
PB 81-117897
APPENDIX A Methodology for Developing Criteria
APPENDIX B Methodology for Developing Criteria
APPENDIX C Methodology for Developing criteria
BIBLIOGRAPHY
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SUMMARY CHART
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11
ACENAPHTHENE
CRITERIA:
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.
(45 F.R. 79318, November 28, 1980}
SEE APPENDIX B FOR METHODOLOGY
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ACROLEIN
CRITERIA^
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.
Human Health
For the protection of human health from the toxic properties
of acrolein ingested through contaminated 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
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13
ACRYLONITRILE
CRITERIA^
Aquatic Life
The available data for aerylonitrile 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
toxicity.
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 10~5, 10~6, and 10~7.
The corresponding recommended criteria are 0.58 ug/L, 0.058
ug/L, and 0.006 ug/L, respectively. If these estimates are made
for consumption of aquatic organisms only, excluding consumption
-------
14
of water, the levels are 6.5 ug/L, 0,65 ug/L, and 0.065 ug/L,
respect ively.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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CRITERIA:
5
AESTHETIC QUALITIES
All waters free from substances attributable
to wastewater or other discharges that:
(1) settle to form objectionable deposits;
(2) float as debris, scum, oil, or other
natter to form nuisances;
(3) produce objectionable color, odor, taste,
or turbidity;
(4) injure or are toxic or produce adverse
physiological responses in humans,
animals or plants? and,
(5) 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.
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16
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 MATER, JULY 1976) FB-263943
SEE APPENDIX C FOR METHODOLOGY
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17
CRITERION:
20 mg/L or more as CaCQ3 freshwater aquatic life except where
natural concentrations are less.
INTRODUCTION:
Alkalinity is the sura 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
alkalinity. Alkalinity resulting from naturally occurring
-------
18
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 (NATO, 1968).
-------
13
Excessive alkalinity can cause problems for swimmers by
altering the pH of the lacritnal fluid around the eye, causing
irritation.
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
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 as 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
-------
so
TABLE I*
Maximum Alkalinity In Waters Used As A Source
Of Supply Prior To Treatment
Alkalinity
Industry mg/L as
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
Bottled and canned soft drinks. ..... ...... 85
* 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
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22
*ALDRIH-DIELDRIN
CRITERIA;
Aguatic Life
pieldrin
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 al.drin
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
* 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 10~5, 10~6 and 10"7
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.
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 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 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.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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O 1
i"—' i
AMMONIA
SUMMARY;
All concentrations used herein are expressed as un-ionized
ammonia (NH3), because NH3, not the ammonium ion (NH^+) 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 NH3 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
0.14 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 un-
ionized 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
-------
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 and pH, 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
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" fishesj
rather, effects of organism sensitivities on the criterion are
most appropriately handled by site-specific criteria derivation
procedures.
Data for concentrations of NH-j 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 Macrobrachivain rosenberg^i. 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 toxicityr (5) more histopathologica1 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 rog/L NH-,) 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-TCAPS; TCAP < T < 30
10-0.03(20-1); 0 <_ T < TCAP
FPH =1 ; 8 < pH < 9
1+10-7.4-PH
1.25 ; 6.5 < pH < 8
1CAP = 20 C; Salmonids or other sensitive
coldwater species present
= 25 C; Salmonids and other sensitive
coldwater species absent
(*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 un-ionized ammonia
(in mg/L NH-j) does not exceed, more often than once every 3 years
on the average, the average* numerical value given by
0.8O/FT/FPH/RATIO, where FT and FPH are as above and;
-------
'**"*
(**J
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
= 20 C; Salmonids and other sensitive
coldwater species absent
(*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 un-
ionized 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
-------
30
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 site-
specific 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
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 1Q5 or 1Q1O for
Criterion Maximum Concentration design flow and 7Q5 or 7Q10 for
the Criterion Continuous Concentration design flow in steady-
state models for unstressed and stressed systems respectively.
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31
(!) On*-hiour ftmrmgm conctnrratloni tor
PM
A. Sal"
0 C
jntds or 0+tw
3 C
10 C
SwisitN* CotdMtw
15 C
So*clM Pr
Un- fool x«0 Am*onlt {ntg/il
6.30
6.7S
7.00
7.23
7*71
8.00
8.25
8.50
8.79
9.00
6.90
6.75
7.00
7.25
7.50
7.75
8.00
8.29
8.30
8.75
9.00
0.0091
0.0149
0.023
0.034
0.049
0.05ft
0.009
0.003
0.009
0.009
0.005
35
32
28
23
17.4
12.2
80)
4.3
2.6
1.47
0.86
0.0129
0.021
0.033
0.048
0.004
0.080
00)92
0.092
0.092
0.092
0.092
Total
33
30
26
22
16.3
11.4
7.3
4.2
2.4
1.40
0.83
0.0182
0.030
0,046
0.068
0.091
0.113
0.130
0.130
0.130
0.130
0.130
Ammonia
31
28
23
20
15.5
10.9
7.1
4.1
2J
t.37
0.83
i. Salmon Ida and Qtnw Swtiltlv* Col
-------
30
f-t
12) 4-day a*»rag« concentrations for awnonla.
OH
0 C
A. Salmon Ids or QtHer
6.50
6.75
7.00
7.25
7.50
7.75
8.00
8,25
8,50
8.75
9.00
6,50
6.75
7.00
7.25
7.50
7.75
6.00
8.25
a, so
8.75
9.00
0.0007
0.0012
0.0021
0.0037
0.0066
0,0109
0.0126
0.0126
0.0126
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
5 C 10 C 15 C 20 C
Sensitive Cold»iter SD»cl»s Present
On-Ionized Ammonia tag/liter NHjJ
0,0009 0.0013 0.0019 0.0019
0,0017 0.0023 0.0033 0.0033
0.0029 0.00*2 0.0059 0.0059
0,0052 0.0074 0.0105 0.0105
0.0093 0.0132 0.0186 0.0186
O.OT53 0.022 0.03 i 0.031
0,0177 0.025 0.035 0.035
0.0177 0,025 0.035 0.035
0.0177 0.025 0.035 0,035
0,0177 0.025 0.035 0.035
0,0177 0.025 0.035 0.035
Total Ammonia (mg/ liter NHj)
2,4 2.2 2.2 1,49
2.4 2.2 2.2 1.49
2,4 2.2 22 1.49
2.4 2.2 2.2 1.50
2.4 2.2 2.2 1.50
2.2 2.1 2.0 t.40
1.44 1.37 1.33 0.93
0.82 0.78 0.76 0.54
0.47 0.45 0.44 0.32
0.27 0.26 0.27 0.19
0.16 0.16 0.16 0.13
i5 C
0.0019
0,0033
0.0059
0,0105
0.0186
0,031
0.035
0.035
0.035
0.035
0.035
.04
.04
.04
.04
.05
0.99
0.66
0.39
0.23
0.15
0.10
30 C
0.0019
0.0033
0.0059
o.otos
0.0186
0.031
0.035
0.035
0.035
0.035
0,035
0.73
0.73
0.74
0.74
0.74
0.71
0.47
0.28
0.17
0.11
0.08
6, Salmon) ds 'and Other Sensitive Cold»atar Species Absent
6.50
6.75
7.00
7.25
7.50
7.75
8.00
a. 25
a. 50
8.75
9.00
6.50
6.75
7.00
7.25
7.50
7.75
8.00
8,25
8.50
8.75
9.00
0 ,0007
0.0012
0,0021
0.0037
0.0066
0.0109
0.0126
0.0126
0,0126
0.0126
0.0126
2.5
2.5
2.5
2.5
2.5
2.3
1.53
0.87
0.49
0.2B
0.16
Un-lonlzed Ammonia dug/liter NHj)
0.0009 0,0013 0.0019 0,0026
O.OOJ7 0.0023 0.0033 0.0047
0,0029 0.0042 0,0059 0.0083
0.0052 0,0074 0.0105 0.0148
0.0093 0.0132 0.0186 0.026
0.0153 0.022 0.031 0.043
0.0177 0.025 0.035 0.050
0.0177 0,025 0.035 0,050
0.0177 0.025 0.035 0.050
0.0177 0.025 0.035 0.050
0.0177 0.025 0.035 0.050
Total Ammonia (mg/llter NHj)
2,4 2.2 22 2.1
2.4 2.2 2.2 2.1
2.4 23 2.2 2.1
2.4 2.2 2.2 2.1
2.4 2.2 2.2 2.1
2.2 2.1 2.0 1.9B
1.44 1.37 1.33 1.31
0.8Z 0.78 0.76 0.76
0.47 0,45 0.44 0,45
0.27 0.26 0.27 0.27
0.16 0.16 0,16 0.17
,
0.0026
0.0047
0.0083
0.01*8
0.026
0.043
0.050
0.050
0.050
0.050
0.050
1.4fi
1.47
1.47
1.48
t.49
1.39
0.93
0.54
0.33
0.2!
O.t4
0.0026
0.0047
0.0083
0.0(48
0.026
0.043
0.050
0.050
0.050
0.050
0.090
.03
,04
.04
.05
.06
.00
0.67
0.40
0.2J
0.16
0.11
To convert these values to mg/llter N, multiply by 0,822.
SIt»-ipeel f Ic criteria development Is strongly suggested at temperature* above 20 C
because of the limited data available to generate the criteria reeoHMndctlon, aivJ
it temperatures belov 20 C because of the limited data and becatlM SM| I change* in
the criteria nay have significant impact on the level of treatment required In
meeting the recommended criteria.
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33
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
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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
ug/L.
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
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35
ARSENIC
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 arsenic apparently
vary from species to species. For inorganic arsenic(III) 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 acute-
chronic 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,
The available data indicate that freshwater plants differ a
great deal as to their sensitivity to arsenic(III) and
arsenic(V). in comparable tests, the alga, Se_l en_ast_rum
capricormitum, 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(III) were in the same range as
the available chronic values for freshwater animals; several
-------
3G
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(III) 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(III) with embryos and larvae of a toad.
Twelve species of saltwater animals have acute values for
inorganic arsenic(III) 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 ..III) 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(III).
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(lll)
ioes not exceed 190 ug/L more than once every 3 years on the
-------
37
average and if the 1-hour average concentration does not exceed
360 ug/L wore 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(III) 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 Skel_etonejma cosrjarum or
Thaiassiosira aestiyalis 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(III) 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
-------
3S
plants at 13 to 56 ug/L. No data are available concerning the
chronic toxicity of any form of arsenic other than inorganic
arsenic(III) to saltwater aquatic life.
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.
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(III) 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
-------
39
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 1Q1O for
Criterion Maximum Concentration design flow and 7Q5 or 7Q1O for
the Criterion Continuous Concentration design flow in steady-
state models for unstressed and stressed systens 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;
For the maximum protection of human health from the potential
carcinogenic effects due to exposure of arsenic 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 10~6, and 10~7. 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
calculated by use of the Guidelines. The risk estimate range is
presented for information purpoes and does not represent an
Agency judgment on an "acceptable" risk level.
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40
(45 F.R. 79318 Nov. 28,1980) (50 F.R. 30784, July 29, 1985)
SEE APPENDIX A FOR METHODOLOGY
-------
.11
ASBESTOS
CRITERIA:
Aquatic Life
No freshwater organisms have been tested with any asbestifonn
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 10 ,
10~6, and 10~7 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 METHODOLOGY
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BACTERIA
CRITERIA
Freshwater Bathing
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!W
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 Bathing
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.
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13
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 MPN per 100 ml for the taking of shel If ish,
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. ^
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44
The quantitative relationships between the rates of swimming-
associated 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 nonswiminers from
that for swimmers. These 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
subtracting 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 large enough 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
"by grouped trial days" approaches.
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The methods used to enumerate the bacterial indicator
densities which showed the best relationship to swimming-
associated gastroenteritis rates were specifically developed for
the recreational water quality studies.
These membrane filter methods have successfully undergone
precision and bias testing by the EPA Environmental Monitoring
and Support Laboratory. ^
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 200 per 100 ml, recommended in 2B§.iitZ Criteria for Water
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 Dufour^ ' 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 noncom-
pliance 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
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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
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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 MPNf 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 shellfish 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 shellfish-
growing 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
show that a 70 coliform MPN per 100 ml at the 50th percentile was
equivalent to a fecal coliform MPN of 14 per 1UO ml. The data,
therefore, indicate that a median value for a fecal coliform
standard is 15 and the 90th percentile should not exceed 43 for a
5-tube, 3-dilution method.
EPA is currently (1986) co-sponsoring, with the National
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49
Oceanic and Atmospheric Administration, research into the
application cc the enterococci and 1. 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
coliform bacterial levels. ^
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CRITERIA FOR INDICAltR FOR BftCTEKIQUDGlCAL OMITIES
Single Sample Maximum Allowable Density (4)j
Acceptable 9*inmlng
Associated Gastro-
enteritis fete per
1000 swlnmers
Freshwater
enterooocci
E_. coli
Marine Hster
enterococci
Notes;
Steady State
Geonetric Mean
Indicator
Density
33<1)
19
35(3)
Designated
Beach Area
(3)
___ ^
9.40
Calculated to nearest whole number using equations
{mean E. coli density) = antilcgjg illness rate/1000 people +• 11.74
g>4Q
Calculated to nearest whole number using equation:
(mean enterococci density) = antilogy illness rate/1000 people - 0.20
12,17
Single sample lImit=antilogiQ ! Logic* indicator geonetzic + Factor detennined Cron x (log^o stard.
mean density/100 ml> areas under the ItoBnal deviation)
probability curve for
the assumed level of
probability
The appropriate factors for the indicated one sided confidence levels arej
75% C.L. - .675
92% C.L. - .935
901 C.L. - 1.28
95% C.L. - 1.65
01
o
(5) Based on the observed log standard deviations during the EEft studies? 0.4 for freshwater E. coli
ant! enterococcif a
standard deviation
ant! enterococcif and 0.7 for marine water enterococci. Each jurisdiction should, establish its own
iation for its conditions s*ich would then vary the single sample limit.
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51
References
1- 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 f: PB 86-158-045)
2* 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)
3* Dufour, A. P. 1983. Health Effects Criteria for Fresh
Recreational Waters. EPA-600/1-84-004, U.S. Environmental
Protection Agency, Cincinnati, OH.
4* 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.
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fro
BARIUM
CRITERION:
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, BaSO4 and
witherite, BaC03, both of which are highly insoluble salts. The
metal is stable in dry air, but readily oxidized by hutnid 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).
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 in a few midwestern and western States ranging from 100 to
3,000 ug/L (PHS, 1962/1963; Katz, 1970; Little, 1971). Barium
enters the body primarily through air and water, since
appreciable amounts are not contained in foods (NAS, 1974).
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53
The fatal dose of barium for nan 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, Ii57). 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
and blood vessels, a limit of 1 mg/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
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54
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 will preclude the
existence of the toxic soluble form under usual marine and fresh
water conditions, a restrictive criterion for aquatic life
appears unwarranted.
(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943
SEE APPENDIX C FOR METHODOLOGY
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BENZENE
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 low as 700 ug/L with a 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 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 6.6 ug/L, 0.66 ug/L, and
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5G
0.066 ug/L, respectively. If these estimates are made for
consumption of aquatic organisms only, excluding consumption of
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
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BENZIDINE
CRITERIA;
Aquatic Life
The available data for benzidine 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 10 ,
10 , and 10~7. 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
ng/L, and 0.05 ng/L, respectively.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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58
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.
Hunan 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 10 , 10 , and 10 . The
corresponding recommended criteria are 37 ng/L, 3.7 ng/L, and
0,37 ng/L, respectively. If these estimates are made for
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59
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
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60
BORON
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 °C 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, I960).
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61
In the dairy cow, 16 to 2O g/day of boric acid for 4O 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
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Grt
f.
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 Dag.hnjLa macjna 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-specific. 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
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G-3
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. Studies with microalgae
and macroalgae revealed effects at 22,8 to 860 ug/L.
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:
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
e{0.7852 [ In(hardness) ]-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 e(i.!28 [ In(hardness) ] -
3.828) more than once every 3 years on the average. For
example, at hardnesses of 50, 100, and 200 mg/L as CaCO3 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
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64
l.B, 3.9 and B.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 "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
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65
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.
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 1Q5 or 1Q1Q for
Criterion Maximum Concentration design flow and 7Q5 or 7Q10 for
the' Criterion Continuous Concentration 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).
HTJMAN HEALTH CRITERIA;
The ambient water quality criterion for cadmium is
recommended to be identical to the existing drinking water
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6G
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 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. 307S4, July 29, 1985)
SEE APPENDIX A FOR METHODOLOGY
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6t~1
{
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
earbontetrachloride 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 10~5, 10~6, and 10~7, The
corresponding recommended criteria are 4.0 ug/L, 0,40ug/L, and
0.04 ug/L, respectively. If these estimates are made for
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68
consumption of aquatic organisms only, excluding consumption of
water, the levels are 69.4 ug/L, 6,94 ug/L, and 0.69 ug/L
respectively.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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69
CHLORDANE
CRITERIA i
Aquatic Life
For chlordane the criterion to protect freshwater aquatic 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
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 10~5, 10~6, and 10~7. 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
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70
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 1.5 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
Monochlorobenz ene
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
potential adverse human health effects.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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CHLORINATED ETHANES
CRITERIA:
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 1,2-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 1,1,2-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
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 1,2-dichloroethane,
31,200 ug/L for 1,1,1-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 10~5, 10~6, and 10~7. 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 1,1,2-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 1Q~5, 10~6, and 10~7.
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.
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73
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 10""5, 1Q~6, and 1Q~7.
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.
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 10~5, 10~6, and 10~7. 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/L, 8.74 ug/L, and 0.87 ug/L,
respectively.
-------
For the protection of human health from the toxic properties
of 1,1,1-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 1,1,1-trichloroethane ingested through contaminated
aquatic organisms alone, the ambient water criterion is
determined to be 1,03 ug/1.
Because of insufficient available data for monochloroethane,
1,1-di ch 1 oroethane , 1,1,1,2-tetrach1oroethane, 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
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75
CHLORINATED NAPHTHALENES
CRITERIA:
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.
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
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76
CHLORINE
SUMMARY;
Thirty-three freshwater species in 28 genera have been
exposed to TRC and the acute values range from 28 ug/L for
5§.£lllli§ HUSO!3, 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, Men_id_ia genjlnsu^ae, anc* i° 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
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t f
unacceptably if the 4-day average concentration of total residual
chlorine does not exceed 11 ug/L more than once every 3 years on
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 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 1Q1O for
Criterion Maximum Concentration design flow and 7Q5 or 7Q10 for
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the Criterion Continuous Concentration 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).
(50 F.R. 30784, July 29, 1985)
SEE APPENDIX A FOR METHODOLOGY
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CHLORINATED PHENOLS
CRITERIA:
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 recogni2ed that
organoleptic data have limitations as a basis for establishing a
water quality criterion, and have no demonstrated
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so
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
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81
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 effects.
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-trichlorophenol
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/Lr
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-raethyl-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
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83
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 quality criterion, 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
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84
CjjLOROAIJCYL ETHERS
CRITERIA;
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 available concerning the chronic toxicity of
chloroalkyl ethers to sensitive freshwater aquatic life,
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-chloroisopropyl) 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-chloroisopropyl) 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(chloromethyl) 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
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85
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, 1Q~6, and 10~7.
The corresponding recommended criteria are 37.6 x 10~6
ug/L, 3.76 x 10~6 ug/L, and 0.376 x 10~6 ug/L,
respectively. If these estimates are made for consumption of
aquatic organisms only, excluding consumption of water, the
levels are 18.4 x 10~3 ug/L, 1.84 x 10~3 ug/L, and 0.114 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 10~5, 1Q~6, 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
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CHLOROFQgM
CRITERIA:
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
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 arobient 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,
10~6, and 10~7. The corresponding recommended criteria are
1.90 ug/L, 0.19 ug/L, and 0.019 ug/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
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88
CHLOROTHENOXY HERBICIDES
2,4-D^ 2t4_f_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 no-
effect 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 for disseminated
coccidioidoinycosis, the patient had no side effects from 18
intravenous doses during 33 days; each of the last 12 doses in
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the series was 800 mg (about 15 mg/kg) or more, the last being
2000 rag (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
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so
CHROMIUM (VI)
AQUATIC LIFE SUMMARY;
Acute toxicity values for chromium(VI) 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(VI5
decreases as hardness and pH increase.
The chronic value for both rainbow trout and brook trout is
264.6 ug/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(VI). 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.
The acute toxicity of chromium (VI) to 23 saltwater vertebrate
and invertebrate species ranges from 2,000 ug/L for a polychaete
worm and a mysid to 105,000 ug/L for the mud snail. The chronic
values for 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(VI)
range from 125 to 236 for bivalve molluscs and polychaetes.
CHROMIUM (III)
Acute values for chromium (III) are available for 20 freshwater
animal species in 18 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(III) being more toxic in
soft water.
A life-cycle test with Dap_h_n_ia magna 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_h_n_ia magna, 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
watermi1foi1. A freshwater green alga was affected by a
concentration of 397 ug/L in soft water. No bioconcentration
factor has been measured for chromium(III) with freshwater
organisms.
Only two acute values are available for chromium (III) in
saltwater 10,300 ug/L for the eastern oyster and 31,500 ug/L for
the muminichog. 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
-------
a4"*
organisms and chromium(lll) range from 86 to 153, similar to the
bioconcentration factors for chromium(VI) and saltwater species.
NATIONAL CRITERIA:
CHROMIUM(vT)
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(VI)
does not exceed 11 ug/L more than once every 3 years on the
average and if the 1-hour average concentration does not exceed
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(VI) 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(III)
The procedures described in the Guidelines indicate that,
except possibly where a locally important species is very
sensitive, freshwater aquatic organisms and their uses should not
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Q'
be affected unacceptably if the 4-day average concentration (in
ug/L) of chromium(III) does not exceed the numerical value given
by e(0,8190[ln(hardness)]-H.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
e(0.8190[In(hardness) j+3.688) more than once every 3 years on the
average. For example, at hardnesses of 50, 100, and 200 mg/L as
CaCOj the 4-day average concentrations of chromium(lll) 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 not affect a polychaete worm in a life-cycle
test.
EPA believes that a measurement such as "acid-soluble" would
provide a more 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 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 exceedenee 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 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 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 1Q5 or 1Q1O for
Criterion Maximum concentration design flow and 7Q5 or 7Q1O or
the Criterion Continuous Concentration 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;
For the protection of human health from the toxic properties
of Chromium III ingested through water and contaminated aquatic
-------
organisms, the ambient water criterion is determined to be 170
mg/L.
For the protection of human health from the toxic properties
of Chromium III 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 grains
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
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3G
2-CHLOROPHENOL
CRITERIA:
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-
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 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.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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e/1
COLOR
CRITERIA:
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
increased color (in combination with turbidity) 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:
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
-------
3
8
color "units" 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 (I963a)
characterized color-causing colloids examined as aromatic,
polyhydroxy, methoxy carboxylic acids. Shapiro (1964)
characterized color-causing constituents as being dialyzable and
composed of aliphatic, polyhydroxyl carboxylic 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
color" 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
-------
oq
..»«. I
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.
The effects of color 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
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
standard or blue end of the spectrum (Birge and Juday, 1930). In
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100
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 360
Chemical and allied products 500
Petroleum 25
-------
101
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.
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 effective photosynthetic oxygen
production. To maintain satisfactory biological conditions, this
depth cannot be substantially reduced.
Industrial requirements as related to water color have been
standardized (HAS, 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
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102
*COPPER
AQUATIC LIFE SUMMARY:
Acute toxicity data are available for species in 41 genera of
freshwater animals. At 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 for 15 freshwater species and
range from 3.873 ug/L for brook trout to 60.36 ug/L 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 suspended, canceled or restricted by U.S.EPA Office
of Pesticides and Toxic substances
-------
103
saltwater algal species have been tested, and effects were
observed between 5 and 100 ug/L. Oysters can bioaccuniulate
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.
NATIOHAL 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 e(o.8545 [ In(hardness) ]-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 e(0.9422 [ In(hardness) ] -1.464) more than once every
3 years on the average. For example, at hardnesses of 50, 100,
and 200 mg/L as CaCQg 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 "acid-soluble" would
-------
104
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 1Q1O for
-------
105
Criterion Maximum Concentration design flow and 7Q5 or 7Q1O 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 rog/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.
(45 F.R. 79318 Nov. 28,1980} (50 F.R. 30784, July 29, 1985)
SEE APPENDIX A FOR METHODOLOGY
-------
10G
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, 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 'ceed 5.2 ug/L more than once every 3 years on the average
and -f 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
-------
108
recover differ greatly, however, and site-specific criteria
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 nay make their use
impractical, in which case one should rely on a steady-state
model. The Agency recommends the interim use of 1Q5 or 1Q10 for
Criterion Maximum Concentration design flow and 7Q5 or 7Q10 for
the Criterion Continuous Concentration 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
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.
-------
109
DOT AND METABOLITES
CRITERIA;
Aquatic Life
DDT
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 o.ooio
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
-------
110
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 10~s, 10~6 and 10~7. The
corresponding recommended criteria are 0.24 ng/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
-------
Ill
DEMgTON
CRITERION;
0.1 ug/L for freshwater and marine aquatic life
RATIONALE:
Static LC50 bioassays yielded toxicity values for the organo-
phosphorus 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). Hesults of these tests demonstrate
an apparent sharp division in species sensitivity, with bluegill
(Lepomis ina.croch_irus_), rainbow trout _L§_a.l!E° gairdneri) and
guppy, IPoecj.^j.a ret_icu_l ataj_, 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 for
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 4S 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 (Notemiggnus
££Ms.°ie.M£§Sl' 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 Ch_ironoinus P_,lujl£sus
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 Gammaru_s f_asc.iatus (Sanders, 1972). These 24- and
96-hour LC50 values are reported as 500 and 27 ug/L, indicating a
time-related effect not observed in the bioassays with fishes.
As only a few of the sensitive species have been tested and great
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
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113
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 AChE-
inhibiting 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 enzyme-
inhibiting 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
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114
PICHLOROB1NZENES
CRITERIA:
Aquatic Life
The available data for dichlorobenzenes indicate that acute
and chronic toxieity 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
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
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pICHLOROBENZIDINE
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 10~5, 10~6, and 10~7.
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.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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11G
DICHLOROETHYLENES
CRITERIA;
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 1,1 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*"5, 10-
6, and 10 . The corresponding recommended criteria are
0,33 ug/L, 0.033 ug/L, and 0.003 ug/L, respectively. if
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117
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.
lj 2-Dichloroejthylene
Using the present guidelines, a satisfactory criterion cannot
be derived at this time because of insufficient available data
for 1,2-dichloroethylene.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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118
2,4-DICMLOROPHENOL
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 2B, I960)
SEE APPENDIX B FOR METHODOLOGY
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DICHLOROPROPANES/DICHLOROPROPENES
CRITERIA:
Aquatic Life
The available data for diehloropropanes 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 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.
-------
ISO
Hunan 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
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2,4-DIMETHYLPHENQL
CRITERIA
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-
diinethyl-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
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DIlflTROTOLUENE
CRITERIA;
Aguatic 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.
Hunan 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 10~5, 10~6 and 1Q~"7. 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
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12-1
DIPHENYLHYDRAZINE
CRITERIA;
Aguatic 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 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 , 10 , and 10 .
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,
respectively.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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.25
ENDOSULFAN
CRITERIA:
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
-------
12G
*ENpR_IN
CRITERIA^
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 U.S. EPA Office
of Pesticides and Toxic Substances
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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ETHYLBEHZENE
CRITERIA:
Aquatic Life
The available data for ethylbenzene indicate that 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.
Hunan 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
mg/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 O.l mg/kg/day for ethylbenzene.
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128
FLUORANTHENE
CRITERIA:
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 fluoranthene to
sensitive freshwater aquatic life.
The available data for 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
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, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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129
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 disease".
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
usually at supersaturation levels higher than those lethal to
fish.
The standard method of analyzing for gases in solutions has
been the Van Slyke method (Van SlyJce 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
to water. Gases pass from the water through the tube, thus
raising the internal gas pressure which is measured by a
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130
manometer or pressure gauge connected to the tube (HAS, 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 (N2), 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
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131
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 for 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.
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 at the base of the dam. The momentum
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
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132
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 spilling 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
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133
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 40 to 95 percent do occur and a major portion
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 tshawvtsena, and steelhead trout,
Sal mo gairdneri, 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),
Bucker (1974), using juvenile coho salmon, Oneo. rhy_nehus
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
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134
saturation even though total gas saturation remained at 119
percent saturation, indicating the importance of determining the
concentration of the individual components (O2 and N2) of the
atmospheric supersaturation. Nebeker et al, (I976a), using
juvenile sockeye salmon, Oncorhv_n£hus 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 CQ2
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 gas-
supersaturated water followed by sockeye salmon, Oncorhyncnus
nerk_a. Chinook salmon, 2!l£°EJlXB£Jlii§ J=shawv_tscha , were
intermediate in sensitivity. echo salmon, Oncorhyncnus kisutch,
were significantly the more tolerant of the salmonids though
still much more susceptible than non-salmonids like bass or carp.
UHEHHi3. 113.2113. exhibited a sensitivity to supersaturation
similar to that of the salinonids {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.
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135
No differences are proposed in the criteria for freshwater
and marine aguatic 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, Brevoortia
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 off 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
105 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 in the fins and occasionally in the eye. At 120 percent
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
organs. The bulbous arteriosis and swim bladder were severely
distended, and emboli were found along the length of the gill
arterioles, resulting in hemostasis. At water temperatures of 30
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136
°C, menhaden did not survive, regardless of gas saturation level.
At water temperatures of 15 , 22 , and 25 °C 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
°C) Menhaden survival after 96 hours at 110 percent nitrogen
saturation ranged from 92 percent at 22° and 25° to 83 percent
at 15 °C. 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 BoucJc 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. I976b).
(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943
SEE APPENDIX C FOR METHODOLOGY
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137
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, 196i; 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, Pimephales
p_rowe]^as_, exposed during a complete life cycle. An estimated
"safe" long-terra 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.
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 5O percent loss of guthion in about 48
hours. Flint et al. (1970) determined the half-life of guthion
at 3OC in pond water and in a phosphate buffer protected fron
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
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13S
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, L§E°™iE El^S-^LtliS.3.' 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, CYE£iHodon
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139
yariegatus, 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 rag/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,
kad_i aken^is , 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, Ga_mma.rus f a s ci. a. t u s , 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 pacif ica
of 2.Q and 0.24 ug/L, respectively, whereas for Pteronarcy_s
californica the values were 4.6 and 1.3 ug/L, respectively.
Results of other toxicity studies on marine organisms have
been reported. The 24-hour LC50 for the white mullet, Mu_g_il_
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140
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 liaanda, to
be 10 to 30 ug/L. The 48-hour LC50 for the European shrimp,
Cranggn crangon, was found to be 0.33 ug/L guthion (Portman,
1972). Butler (1963) found that the 24-hour EC50 for blue crab,
s a pi dug, was 550 ug/L and the 48-hour EC5Q for pink
shrimp, Penaeus duorarum, was 4.4 ug/L guthion. The 4 8 -hour TLa
was estimated to be 620 ug/L for fertilized oyster eggs,
Crassostrea vjjrejijvica., 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 9 6 -hour LC50 of 0.1
ug/L for Gajjmarus 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
-------
Ill
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 haloethers.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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142
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 chronic 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.
Human Health
For the maximum protection of human health from the potential
carcinogenic effects of exposure to chlororoethane,
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 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
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1-13
are estimated at 10~5, 10~6 and 10~7. The corresponding
recommended criteria are 1.9 ug/L, 0.19 ug/L, and 0.019 ug/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
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144
HARDNESS
_INTOODUCTION;
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 (CaCC^).
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
Cone. mg/L CaCO3 Description
0-75 soft
75 - 150 moderately hard
150 - 30O hard
300 and up 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.
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115
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)
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116
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 (MAS,
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 (HAS, 1974). Subsequent
treatment generally can reduce hardness to tolerable limits
although costs of such treatment are an important factor in
determining its desirability for 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
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147
TABLE 4.
Maximum Hardness Levels Accepted
By Industry as a Raw Water Source*
Maximum Concentration
Industry SS£L as CaC03
Electric utilities 5,000
Textile 12O
Pulp and paper 475
Chemical 1,000
Petroleum 900
Primary metals 1,000
* Requirements for final use within a process may be essentially
zero, which requires treatment for concentration reductions.
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1-1S
than hardness.
(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943
SEE APPENDIX C FOR METHODOLOGY
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119
HEPTACHLOR
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.
Hunan 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 10~5, 10~6, and 10~7. 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,
respect ively.
(45 F.R. 79318, November 28, 1980}
SEE APPENDIX B FOR METHODOLOGY
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150
HEStACHLDROBUTADIENE
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 32 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 hexachlorobutadiene
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 10~5, 10 , and 10~7. The
corresponding recommended criteria are 4.47 ug/L, 0.45
ug/L, and 0.045 ug/L, respectively. If these estimates are
made for consumption of aquatic organisms only, excluding
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151
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
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HEXACHLOROCYCLOHEXANE
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 availabTe
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.
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153
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 10~5, 10-6^ and 1Q-
7. The corresponding recommended criteria are 22 ng/L, 2.2
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.
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 10~5, 10~6, and 10~7.
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.
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151
For the maximum protection of human health from the potential
carcinogenic effects due to exposure of r-
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 10~5, 10~6, and 10~7. 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 technical-
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 1Q"~5, 10"6, and 10~7. 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.
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155
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
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156
HEXACHLOgOCYCLOPENTADIENE
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
hexachlorocyclopentadiene to sensitive saltwater aquatic life.
Hunan 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
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IRON
CRITERIA;
0,3 mg/L for domestic water supplies (welfare).
l.O rog/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"1""1") , and the ferric, or trivalent
p
*• ) 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 organonietal 1 ic 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
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15S
life.
(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943
SEE APPENDIX C FOR METHODOLOGY
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159
ISOPHORONE
CRITERIA:
Aguatic 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.
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
mg/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)
SEE APPENDIX B FOR METHODOLOGY
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160
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.
Acute values are available for 13 saltwater animal species
and range from 315 ug/L for the muinmichog to 27,000 ug/L for
the soft shell clam. A chronic toxicity test was conducted
with a mysld; 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:
The procedures described in the Guidelines for Deriving
Numerical National Water Quality Criteria for the Protection of
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161
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
e(l»273 [ In(hardness) ]-4.705) 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 e (1.273[In(hardness)]-
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 "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
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162
method. This has two impacts: (I) 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 1Q5 or 1Q1O for
Criterion Maximum Concentration design flow and 7Q5 or 7Q1O for
the Criterion Continuous Concentration 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).
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163
HOMAH 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 wnic 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
exposure so ley from consumption of 6,5 grains 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
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164
MALATHION
CRITERION;
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 gairdneri, 68 ug/L
(Cope, 1965); largemouth bass, Mi.crop_terus_ sa_lmo.ides, 50 ug/L
(Pickering et al, 1962); and chinook salmon, 2.B£££!!Y!!c_hus
t.s.ha.wv^Escha^ 23 ug/L (Katz, 1961). All of the above tests were
in static systems. Eaton (1970) determined a 96-hour LC50 for
bluegill, Ljegomj.s_ ma^croch^rus, in a flow-through system at 110
ug/L. Macek and McAllister (1970) reported a similar 96-hour
LC50 for the bluegill in a static exposure. Static 96-hour LCSOs
of 120 and 160 ug/L were reported by Post and Schroeder (1971)
for brook trout, Salyelinus fontina1is. Bender (1969) indicated
that the acute toxicity to fathead minnows, P_imep_hales prom el as,
is slightly greater (about 2.0 times) in a static system than in
a flow-through system. The flow-through acute toxicity to
fathead minnows reported by Mount and Stephen (1967) approximated
the static acute toxicity reported by Henderson and Pickering
(1958) and Bender (1969).
-------
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 Pteronarcel^a ba^dj-a, 1.1 ug/L
(Sanders and Cope, 1968) ; and for Gammajrus f^s_ci.atus, 0.76 ug/L
(Sanders, 1972), The 48-hour LC50 for Siroocephalus serrulatus
was 3.5 ug/L and for Dapjin^ia EuJ^ex, 1.8 ug/L (Sanders and Cope,
1966). M§.EillliS were immobilized in 50 hours in 0.9 ug/L
(Anderson, 1960). The 24-hour LC50s for two species of roidge
larvae were 2.1 ug/L (Mulla and Khasawinah, 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 sensitive 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
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16G
marine fish, Lagpdon 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, Mugj.1 ceghalus, 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, Le_ios_t£inus_ xanthurus_, and Atlantic croaker,
Micropogon undulatus, also had substantial inhibition of brain
during the spray operation (70 percent or more inhibition).
Toxicity studies have been made on a number of marine
animals. Eisler (1970) studied the 96-hour LC50 for several
marine fishes at 20 °C in static, aerated seawater. The 96-hour
LC5O values (in ug/L) were: Menidja menidia, 125; Mugj.1 cephalus,
550'" Zundu^ujs S^iEii3.' 250; ZHH^MiMf. !!J-t®£££:!iiii.s.» 240;
§E!ia.§£°A£le.s. 5li.cu_latus, 3250; AnguJLiLILa £os>trata, 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,
-------
1SE2E® £5.H§.tiiiS' i" 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, Pa1aemonetes vulgaris?
and 83 ug/L for hermit crab, Pagurus longicarpus. Growth of
oyster, Cr_as_sqstrea 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 AChl 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 Datphnia, 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
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16S
MANGANESE
CRITERIA;
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 (Mn02), 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 form 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 have sufficient manganese, but beef cattle on a
high corn diet may require a supplement.
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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 ISO 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 nay 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 qualities.
McKee and Wolf (1963) summarized data on tojcicity 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 have been reported to kill fish in 8 to 18 hours at
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170
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 (HAS, 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 i' water consumed by
livestock. At concentrations of slightly less than 1 jng/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 are
cultivated and irrigated, a criterion of 200 ug/L is suggested
for consideration.
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171
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) PB-263943
SEE APPENDIX C FOR METHODOLOGY
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±72
*MERCUR¥
AQUATIC LIFE SUMMARY:
Data are available on the acute toxicity of mercury(ll) 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(II).
Available chronic data indicate that methyImercury 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(II) the chronic value
obtained with jDapJmia raag_na 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.
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(II) 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.
*Indicates suspended, canceled or restricted by U.S. EPA
Office of Pesticides and Toxic Substances
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173
Data on the acute toxicity of mercuric chloride are available
for 29 genera of saltwater animals, including annelids, molluscs,
crustaceans, echinoderxns, 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(II).
Results of a life-cycle test with the mysid show that mercury(II)
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(II).
Thus mercury(II) should be the only important possible cause of
acute toxicity and the Criterion Maximum Concentrations can be
based on the acute values for mercury(II).
The best available data concerning long-term exposure of fish
to mercury(II) 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
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17-1.
to exceed 1.0 mg/kg. Although it is not known what percent of
the mercury in the fish was methy 1 mercury, 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 methyImercury, 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(II).
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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 raethyImercury 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
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
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176
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.
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.
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-specific criteria may
be established if adequate justification is provided.
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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 1Q5 or 1Q10 for
Criterion Maximum Concentration design flow and 7Q5 or 7Q10 for
the Criterion Continuous Concentration 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
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
ng/L.
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
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f
METHOXYCHLOR
CRITERIA;
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 rag/kg of body weight/day (Lehman,
1965), Where adeguate human data are available for corroboration
of the animal results, the total "safe" drinking water intake
level is assumed to be 1/100 of the no-effect or minimal effect
level reported for the roost 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 (Pimephales
promelas) 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 96-hour LC50 concentration
was 7.5 and 22 ug/L for the fathead minnow and yellow perch,
respectively. Korn and Earnest (1974) obtained a 96-hour LC50
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179
of 3.3 ug/L with juvenile stripped bass, Morone_ 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, Orconec_tes na_is_. Merna and Eisele (1973)
obtained a 96-hour LC50 value of 0,61 ug/L for the scud, Gammarus
pseudolimnaeus and 96-hour LCSO's ranging from 1.59 to 7.05 ug/L.
for the crayfish, Orconecte_s naJLs, and three aquatic insect
larvae. In 28-day exposures, reduction in emergence of mayflies,
Sjtenoneroa sp., and in pupation of caddisflies, 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, Hyal lei la 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, Grconectes naij;. 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
bioaccutnlative 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
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ISO
tissues have been established by the U.S. Food and Drug
Administration.
The above data indicate that 0,1 ug/L inethoxychlor would be
just below chronic effect level for the fathead minnow and one-
fifth the acute toxicity level in a crayfish species. Therefore,
a criterion level of 0-Q3 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 Nimrao (1974) found the 96-hour LC50 of
methoxychlor for the pink shrimp, Pen_ae_u_s 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, Mercenar_ia mercenari-a and My_a f|r e n a r i. a,
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
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181
MIREX
CRITERION:
0.001 ug/L for freshwater and marine aquatic life.
RATIONALE;
Mirex is used to control the imported fire ant Solengp_s_j.s
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 mirexf 14.7 percent soybean oil and 85 percent corncob
grits. The mirex bait often is applied at a rate of 1.4 kg/ha,
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, 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 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
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hay_i, 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 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-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
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183
(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.
In another study, Van Valin et al. (1988) exposed bluegills,
Lep_£nd,^ nacrochiruB, and the goldfish, Carassius auratus, to
mirex by feeding a mirex-treated diet (1, 3, and 5 »g airex 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.
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184
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. 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. Wolfe and Moment (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 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 osse_us, to
0.05 ug/g in the areal fin of the northern pike, Es_ox .luc_ius.
Mirex has never been registered for use in Canada.
Mirex does not appear to be greatly toxic to birds, with
LC50'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 at 3 ug/g with mallards and 13 ug/g with
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
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1S5
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 flow-through experiments, Tagatz et al. (1975) found
reduced survival of Caj.linectes sapidus, Pena_eu_s djjroraruin, and
grass shrimp, Pa.iaeinonet.es P_u2_ip, 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
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186
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 might be considered
"acute" 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
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NAPHTHALENE
CRITERIA:
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 life 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.
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
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NICKEL
CRITERIA;
Aquatic Life
For 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 [In(hardness) J+1.06) as a
24-hour average, and the concentration (in ug/L) should not
exceed the numerical value given by e(0.76[ln (hardness) ]+4. 02)
at any time. For example, at hardnesses of 50, 100, and 200 rng/L
as CaCO3 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.
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
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
mg/L.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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CRITERION:
INTRODUCTION:
189
HITRATES/NITRITES
10 mg/L nitrate nitrogen (N) for
domestic water supply (health),
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 nitrate-
containing soils become anaerobic and the conversion to nitrite,
molecular nitrogen, or nitrous oxide occurs. Ammonium ions may
also be produced in some circumstances.
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
farm-site fertilizer and animal wastes, lawn fertilizer, leachate
-------
ISO
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,
1372). 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.
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,
-------
191
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 ntethemoglobinemia
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 public 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 bottle-
fed 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
-------
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 gairdneri, 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, iSESfii5. roacjroch_ir_us_, 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, M_icrogterus saj.inoi.des, and channel catfish,
Ictalurus punctatus, could be maintained at concentrations up to
400 ttg/L nitrate (90 ing/L nitrate nitrogen) without significant
effect upon their growth and feeding activities.
The 96-hour and 7-day LC50 values for Chinook salmon,
^BES^llYJlSilH3. "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, SaJ.mo 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, OBc-PJi!!Y-n.<±^.u.s_
tshawytscha, suffered a 40 percent mortality within 24 hours at
-------
1S3
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., winnows, Ph°JiimiJ» iS^Zi3-' "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
for mortality to occur at 10 mg/L (Klingler, 1957), and carp,
Cv.pjr.inus^ carpio, when raised in a water reuse system, tolerated
up to 1.8 ag/L nitrite nitrogen (Saeki, 1965).
Gillette, 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, Garobusia affin is. McCoy
(1972) tested the nitrite susceptibility of 13 fish species and
found that logperch, Perc_ina_ cajarodes, were the most sensitive
species tested (mortality at 5 mg/L nitrite nitrogen in less
than 3 hours of exposure) whereas carp, Cyprinus carp_io, and
black bullheads, ictaiMHH5. SlSi^3.' survived 40 mg/L nitrite
nitrogen for a 48-hour exposure period; the common white sucker,
Catostorous cowme^rsoni, and the quillback, CarpJjDdes cyprinus,
survived 100 mg/L for 48 and 36 hours, respectively.
Russo et al. (1974) performed flow-through nitrite bioassays
in hard water (hardness = 199 mg/L caC03? alkalinity - 176 mg/L
CaCO3; pH = 7,i) on rainbow trout, SjaJLnio g_a_ir;dne_rJL, 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
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134
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 rog/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 roost 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.
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
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1 *}
1 » n,
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 that 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)
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
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136
be substituted when it becomes available. The January,
1986, draft Verified Reference Dose document cites an RfD
of .0005 mg/kg/day for nitrobenzene.
-------
NITROraENOLS
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 wore 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.
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
-------
IBS
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
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MITROSAMIMES
CRITERIA;
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 10~5, 10~6, and 10~7. 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
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200
consumption of aquatic organisms only, excluding consunption 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
carcinogenic effects of exposure to H-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 10~5, 10-
6, and 10~7 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 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 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
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201
ng/L, and 58.7 ng/L, respectively.
For the maximum protection of 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 10""5, 10~6, and 10~7.
The corresponding recommended criteria are 160 ng/L, 16 ng/L, and
1.6 ng/Lr 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 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
49,000 ng/L, 4,900 ng/L, and 490 ng/L, 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.
-------
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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CRITERIA:
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.
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 Tamp__ico Maru wreck of
1957 in Baja, California, (Diaz-Piferrer, 1962), and the No. 2
fuel oil spill in West Palmouth, Massachusetts, in Ii69
-------
204
(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,
-------
205
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 (Blunter, 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 (Bluroer, 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
-------
206
concentrations of 10~4 to 10 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 types 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 organoloptically
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
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.
Oils of animal or vegetable origin generally are chemically
nontoxic to humans or aquatic life; however, floating sheens of
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208
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
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DISSOLVED OXYGEN
NATIONAL CRITERIA;
The national criteria for ambient dissolved oxygen concentra-
tions for the protection of freshwater aquatic life are presented
in Table i. The criteria are derived from the production impair-
ment 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 repre-
sent 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 for 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
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210
salmonids, nonsalmonid coolwater fish, or the sensitive centra-
chid, the smallmouth bass 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 impair-
ment" values given in the document as means and the "slight
production impairment" values as minima. The table which pre-
sents 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 tempera-
tures for the appropriate taxonomic and life stage classi-
-------
Table 1.
211
Water quality criteria for ambient dissolved oxygen
concentration.
'
Coldwater Criteria Warmwater Criteria
30
7
7
1
Day Mean
Day Mean
Day Mean
Minimum
Day
Minimum
Early Life Other Life Early Life
Stages1'2 Stages Stages
NA3 6.5 NA
9.5 (6.5) NA 6.0
NA 5.0 NA
8.0 (5.0) 4.0 5.0
r 5
Other Life
Stages
5.5
NA
4.0
3.0
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.
2
Includes all embryonic and larval stages and all juvenile
forms to 30-days following hatching.
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
l-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 appropriate-
ly 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 calcu-
lated (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.
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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» Nonsalroonid 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
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214
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 (mq/L)
Day Daily Max.
1 9.0
2 10.0
3 11. 0
4 12. Oa
5 10.0
6 11.0
7 12. Oa
1-day Minimum
7-day Mean Minimum
7 -day Mean
Daily Min.
7.0
7.0
8.0
8.0
8.0
9 0
10.0
57.0
7.0
8,1
Daily Mean
8.0
8.5,
9.5b
•9.5
9.0
10. 0
10. 5C
65.0
9.3
a Above air saturation concentration (assumed to be 11.0
mg/L for this example).
b (11.0 + 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 toler-
able so long as:
-------
1. the average concentration attained meets or exceeds the
criterion,'
2. the average dissolved oxygen concentration is calculated as
recommended in Table 3; and
3. 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 roanipulatable discharges.
The previous EPA criterion for dissolved oxygen published in
Quality Criteria^ for Water (USEPA, 1976) was a minimum of 5 rng/L
(usually applied as a 7Q10) which is similar to the current
criterion minimum except for other life stages of waoiwater fish
which now allows a 7-day mean minimum of 4 mg/L. The new
criteria are similar to those contained in the 1968 "Green Book"
of the Federal Water Pollution Control Federation (FWPCA, 1968).
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216
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 labo-
ratory growth experiments and by extrapolation to other activity-
related 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; (l) 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 roost important
factor in general, especially if the acute value is violated. A
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217
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 com-
prised 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 biologi-
cal 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 allocation 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
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218
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 concen-
tration 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 accept-
able 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
-------
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 ng/L for wartnwater fish. Such decisions could be site-
specific based upon the extent of control and serial correlation.
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220
PARATHION
CRITERION;
0.04 ug/L for freshwater and marine aquatic life.
gATIOKALE;
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
inacroc_h_iru^, 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 for bluegills, Le_gom_is
njac_rochj.rus, fathead minnows, Pirosp.haljss Eromel_as_, and brook
trout, £a_3Lvel_iinus fontinalis, respectively. Korn and Earnest
(1974) found a 96-hour LC50 of 18 ug/L for juvenile freshwater
and estuarine striped bass, Mprone saxatilis, in a flowing water
system.
Few chronic exposure data are available for aquatic
organisms. Brown bullheads, Ictalurus 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
that death results from AChE inhibition ranging from 25 to 90
percent of normal. Weiss (1959) also showed that susceptibility
depended upon the extent of recovery of AChE activity following
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, Poeci^^ia reticulata, at 10 ug/L
(Billard and deKinkelin, 1970), alternation of oxygen
consumption rate in bluegills, Lepomis macrqchirus, at 1OO ug/L
(Dowden, 1966), and liver enlargement associated with increased
pesticide-hydrolizing capability in mosquitofish, Gambusia
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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 Dap_hn_ia magnaj 48-hour LC50 of 0.6 ug/L for DjyahjnjLa EHi6.**
48-hour LC50 of O.37 for S_imocepha_:ius_ ^errul^atus (a daphnid)
(Sanders and Cope, 1966); a 5-day LC50 of 0.93 ug/L for the
larval stonefly, Acroneuria pjacif ica (Jensen and Gaufin, 1964);
and a 96-hour LC50 of 0.43 ug/L for the larval caddisfly
SZ5£°.Esv_che cal.i.£orni.£§_ (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 TaHYEH^ girodhaujsi.. Spacie (1975)
obtained 96-hour LCSO's in flow-through bioassays of 0.62 ug/L
for Dap_hn_ia_ amgjia, 0.40 ug/L for the scud, Gajnmarus fasciatus,
and 31.0 ug/L for 4th instar of Ch_ironomous teiitans, 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 LCSO's for Pter^narcyjs £3ii^°£Hi£3 and
^ILSSJrElis?: E^-Si^iE^ of 2-2 ancl 0-44 ug/L, respectively. Spacie
(1975) found the 3-week LC50 for Da_p_hn.ia ma^na 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
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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 "natural" 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, Garobus_ia affinis,
was greater in static than in flowing water test systems.
Sanders (1972), in 96-hour bioassays with the scud, Gammarus
fasciatus, and glass shrimp, £aiaero°!l®t®§. kadlakens^s, also
observed greater toxicity under static than in flow-through
conditions.
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, .Icta^urus inel^as, up to about 50 times water
concentrations. Spacie (1975) found muscle concentrations in
chronically exposed brook trout, S. fontinalis, to be several
hundred times water concentrations; bluegills, LSJDOTOJ.S
IS^^S^IliiMS' na£i about 25 tines water concentrations in their
bodies. Leland (1968) demonstrated a biological half-life of
parathion in rainbow trout, S_a_l2O 2a_iirdjnej:i, 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.
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224
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). Inhibition of brain AChE of fishes by 46 percent or more
has been associated with harmful effects in exposures to
organophosphate pesticides for one life cycle (laton, 1970) and
for short periods (Carter, 1971? Coppage and Duke, 1971 j Coppage,
1972} Coppage and Matthews, 1974? Post and Leasure, 1974; Coppage
et al. 1975). It has been shown that a concentration of 10 ug
parathion/L of flowing seawater kills 40 to 60 percent of the
marine fishes Lagodon rhombo^des (pinfish) and Leostorous
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 yariegatus, 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.
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Other estimates of parathion toxicity to marine organisms
follow. The 48-hour EC50 for parathion to Penaeus duoraruin was
found to be O.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
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PENTACHLOROPHENOL
CRITERIA:
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
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)
SEE APPENDIX B FOR METHODOLOGY
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EH
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 normallyoccurring
range.}
INTRODUCTION;
"pH11 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
(CO2), carbonic acid, (H2C03), bicarbonate ion (HCO.j) and
carbonate ions (C03). 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 (H2S) (Jones, 1964),
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228
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 controlled by the
amounts of alkalinity and acidity present.
RATIONALE:
Knowledge of pH in the raw water used for 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 (HOCl) 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
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 pH 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
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230
to be directly toxic, or to cause the pH range 5 - 6 to
become lethal.
Mount (1973) performed bioassays on the fathead minnow,
Pimephales promelas, for a 13-month, 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 CC>2 is greater than 20 ppm, or
the water contains iron salts which are precipitated as
ferric hydroxide, the toxicity of which is not known.
6.0 - 6.5 Unlikely to be harmful to fish unless free carbon dioxide
is present in excess of 10O ppm.
6.5 - 9.0 Harmless to fish, although the toxicity of other poisons
may be affected by changes within this range.
EIFAC, 1969
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331
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) stoneflies (four species), dragonflies (two
species), and mayflies (one species). All are important fish
food organisms. The 30-day TL5Q 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
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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 by moving out of the area.
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
investigators developed a model which illustrates the effects of
variable pH on copper, zinc, cadmium, and lead; small changes in
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S33
pH 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
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234
PHENOL
CRITERIA;
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.
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
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23J
CRITERION::
PHOSPHORUS
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 bioaecumulation in much the same way as mercury.
Phosphorus as phosphate is one of the major nutrients required
for plant nutrition and is essential for life. In excess of a
critical concentration, phosphates stimulate plant 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
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Z3G
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-wining.
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 "P". 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; (i) 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! and (5) the level of discharge from the lake or of
the penstock from the reservoir.
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 for recycling within the faiotic 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
system,
RATIONALE;
Elemental Phosphorus
Isom (1960) reported an LC5O of 0.105 mg/L at 48 hours and
0.025 mg/L at 160 hours for bluegill sunfish, Legom^s
macrochirus, exposed to yellow phosphorus in distilled water at
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23S
26 °C and pH 7. The 125-and 195-hour LCSO's of yellow phosphorus
to Atlantic cod, Gadus mprhua, 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 p* 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 p^ 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 3OO 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.
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239
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.
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210
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
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 are 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
associated with lakes or reservoirs and currently there are more
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241
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: one 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
evident that a portion of that phosphorus that enters a stream or
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.
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2-12
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.
Mean Depth/Hydraulic
Detention Time
Oligotrophic or
Permissible
Loading
Eutrophic
or Critical
Loading
(meters/year) (grams/meter /year) (grams/meter/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
-------
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 cost-
effective limitations may help control introduced pollutants. 3.
Waters may be highly laden with natural silts or colors which
reduce the penetration of sunlight needed for plant
photosynthesis. 4. Some waters morphoinetric features of steep
banks, great depth, and substantial flows contribute to a history
of no plant problems. 5. Waters 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.
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
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PHTHALftTE 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.
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/1.
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245
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/1.
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-ethyIhexy1 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-ethyIhexy1 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
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24G
PQLYCHLORINATED BIPHENYLS
CRITERIA?
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
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
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 10~5, 10*"^, and 10" ,
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
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POLYNUCLEAR AROMATIC HYDROCARBONS
CRITERIA;
Aguatlc Life
The limited freshwater data base available for polyrmclear
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 10~5, 10~6, and 10~7, 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, excluding consumption of water, the levels are 311.0 ng/L,
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2-19
31.1 ng/L, and 3.11 ng/L, respectively.
(45 F.R. 79318, November 28, 1980}
SEE APPENDIX B FOR METHODOLOGY
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250
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
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£51
standard. For this reason a selective criterion based on
exposure solely from consumption of 6.5 grains of aquatic
organisms was not derived.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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SILVER
CRITERIA:
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 [ In(hardness) ]-6.52} at any
time. For example, at hardnesses of 50, 100, and 200 mg/L as
CaCO , 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 toxieity 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 grains of aquatic
organisms was not derived.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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CRITERION;
253
SOLIDS (DISSOLVED) AND SALINITY
25O mg/L for chlorides and sulfates
in domestic water supplies (welfare)
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, I960). 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 cf 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
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£54
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 ing/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 watex 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)
27O 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 of sodium in
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water supplies is costly (HAS, 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
"excellent" while those with 1,283 to 1,333 mg/L dissolved solids
were "unacceptable" depending on the rating system used. A "good"
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, et al. 1955) evaluated the effect of
chlorides on water used for brewing coffee indicated threshold
concentrations for chloride ranging front 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
protect consumers of drinking water.
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256
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 1,750 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 50.50 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 pike-
perch survived 15,000 mg/L, but only the stickleback survived
20,OOO 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 maintaining osmotic pressure levels of
less than that caused by a 15,000 mg/L solution of sodium
chloride.
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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).
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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 detri-
mental effects will usually
be noticed • • 500
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£59
water which can have detri-
mental effects on sensi-
tive crops • 500-1,000
water that may have adverse
effects on many crops and
requires careful manage-
ment 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).
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
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Table 10.-Total Dissolved Solids Concentrations of Surface
Waters That Have Been Used as Sources for
Industrial Water Supplies
Industry/Use Maximum Concentration
(rog/L)
Textile 150
Pulp and Paper 1,080
Chemical 2,500
Petroleum 3,500
Primary Metals 1,500
Boiler Make-up 35,000
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261
maximum values accepted by various industries for process
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 limiting
factor for industry,
*ISP = 100 [a + b(SAR)]
1 [a + b(SAR)]
where: a = intercept respresenting experimental
error
(ranges from -0.06 to 0.01)
b =slope of regression line (ranges
from 0.014 to 0.016)
**SAR = sodium adsorption ratio = Na
[0.5(Ca + Mg)]°'5
SAR is expressed as rnilliequivalents
(QUALITY CRITERIA FOR WATER, JULY 1976} PB-263943
SEE APPENDIX C FOR METHODOLOGY
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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 nornt for
aquatic life,
INTRODUCTION:
The term "suspended and settleable solids" 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 "solids" is used in this discussion because of
its more common use in the water pollution control literature.
RATIONALE:
Suspended solids and turbidity are important parameters in
both municipal 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
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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 (HAS, 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,*
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(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 solid concantrations (Gammon, 197O).
Similar effects have been reported downstream from an area which
was intensively logged* Major increases in stream suspended
solids (25 ppm turbidity upstream versus 39O ppm downstream)
caused smothering of bottom invertebrates, reducing organism
density to only 7.3 per square foot versus 2S.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).
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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
pollution-sensitive forms disappear (Mackenthun, 1973). Algae
likewise 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
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marketability of some leafy crops like lettuce, and finally the
adverse effect on irrigation reservoir capacity, delivery canals,
and other distribution equipment (NAS, 1974}.
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
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SULFIDE - HYDROGEN SULFIDE
CRITERION;
2 ug/L undissociated H2S for
fish and other aquatic life, fresh
and marine water.
INTRODUCTION:
Hydrogen sulfide is a soluble, highly poisonous, gaseous
compound having the characteristic 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 rag/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.
When soluble sulfides are added to water they react with
hydrogen ions to form HS or H2S, the proportion of each depending
on the pH. The toxicity of sulfides derives primarily from H^S
rather than from the hydrosulfide (HS~) or sulfide (S=) i°ns.
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
at pH 5 about 99 percent of the sulfide is present as H2S (NAS
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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 iong-term data, both in field situations and under
controlled laboratory conditions, demonstrate hydrogen sulfide
toxicity at lower concentrations.
Colby and Smith (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
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263
first 20 ram 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
47 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 (Catgstomus
commersoni) and Smith (1971), Safe levels in working on walleyes
and fathead minnows, Pimephales propelas, 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, Caras_s_ius auratus,
varied greatly in tolerance to hydrogen sulfide with changes in
temperature. They were more tolerant at low temperatures (6 to
1OO C). Holland, et al. (1960) reported that 1.0 mg/L sulfide
caused 100 percent mortality in 72 hours with pacific salmon.
On the basis of chronic tests evaluating growth and survival,
the safe H2S level for bluegill (Leppmis 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 Pseudol imnaeus and Hexagenia I inibata were 2 and 15 ug/L,
respectively (Oseid and Smith, 1974a, 1974b). Some species
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typical of normally stressed habitats, Asellus spp., were much
more resistant (Oseid and Smith, 1974c).
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.
(QUALITY CRITERIA FOR WATER, JULY 1976) PB-263943
SEI APPENDIX C FOR METHODOLOGY
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TAINTING SUBSTANCES
CRITERION:
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,
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 nputral sulfite
pulping processes, municipal wastewater treatment plants, oily
wastes, refinery wastes, phenolic wastes, and wastes from
slaughterhouses. The list of implicated 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
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Z7Z
of fish flesh.
Shuraway and Palensky 1973) determined estimated threshold
concentrations for 22 organic compounds. The values ranged from
0,4 ug/L (2,4-dichlorophenol) to 95,000 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 pollutant. 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
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TEMPERATURE
CRITERIA;
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 tine):
1, One limit consists of a maximum temperature for short
exposures that is time dependent and is given by the species-
specific equation:
Temperature =(l/b)(log [time . ] -a) - 2_ C
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acclimation temperature -minus 2Pto c when the lower
lethal threshold temperature equals the ambient water
temperature (in some regions this limitation may also be
applicable in summer).
or
b. 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 for 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 site-
specific 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.
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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, providing 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:
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Short-term Maximum
Maximum True Daily Mean*
Sub tropical regions (south of 32.2° C (90° F) 29.4° C (85° F)
Cape Canaveral and Tampa Bay,
Florida, and Hawaii
Cape Hatteras, N.C., to 32.2° C (90° F) 29.4° C (85° F)
Cape Canaveral, Fla.
Long Island (south shore) 30.6° C (87° F} 27.8° C (82° 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. Offstreant
domestic uses and instream recreation are both partially
temperature dependent. Likewise, the life 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
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zr>
catalyst, a depressant, an activator, a restrictor, a stimulator,
a controller, a killer, one of the roost important and most
influential water quality characteristics to life in water."
RATIOWAU:
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°
C. 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 35° 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 coliforin bacteria die away more rapidly in
the environment with increasing temperatures (Ballentine and
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27S
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 5° 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
tine, dosages required for equivalent disinfective effect
increased by as much as a factor of 3 when temperatures were
decreased from 20° C to 10° 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,
1967j 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 35° C and blue-greens predominated above 35° C
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279
(Cairns, 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
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 Krenkel, 1969).
The effects of sublethal temperatures on metabolism,
respiration, behavior, distribution and migration, feeding rate,
growth, and reproduction have been summarized by De Sylva (1969),
Another study has illustrated that 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, I960).
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
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280
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
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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 345° C.
In open waters elevated temperatures may affect periphyton,
benthic invertebrates, and fish, in addition to causing shifts in
algal dominance. Trembley (I960) 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
balanced benthic population structure is approximately 32° 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. line-dependent temperature limitations for survival of
brief exposures to temperature extremes, both upper and lower,
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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
juveniles, and temperatures required for metamorphosis,
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 5O percent mortality
relationship is:
log10 (time (minutes)j = a + b (Temperature (° 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 II-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 II-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
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criterion of 2° 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 "a" and "b" for the thermal resistance equation, and
information of size, life stage, and geographic source of the
particular test species (Appendix II-C, MAS, 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 temperature-
sensitive life function appears to be growth (Coutant, 1972).
Coutant (1972) has suggested that a satisfactory estiaate of a
limiting maximum weekly mean temperature may be an average of the
optimum temperature for growth and the temperature for 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 one-
third of the difference between the ultimate upper incipient
lethal temperature and the optimum temperature {MAS, 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
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28-1
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
Laboratory 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 "no stress" level from cold shocks (NAS,
1974). Figure 2 was developed from available data in the
literature (ERL-Duluth, \976) 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.
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28C
TABLE 11.-Example Calculated Values for
Maximum Weekly Average Temperatures for Growth and Short-Term
Maxima for Survival for Juveniles and
Adults During the Summer
(Centigrade and Fahrenheit).
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
Growth3
20
27
32
19
32
18
30
17
32
28
19
25
29
18
28
28
29
(68)
(81)
(90)
(66)
(90)
(64)
(86)
(63)
(90)
(82)
(66)
(77)
(84)
(64)
(82)
(82)
(84)
Maxima
23
35
24
35
24
25
34
30
24
22
(73)
(95)
(75)
(95)
(75)
(77)
(93)
(86)
(75)
(72)
- Calculated according to the equation (using optimum
temperature for growth)
maximum weekly average temperature for growth = optimum
temperature + 1/3 (ultimate incipient lethal temperature-
optimum temperature.
b - Based on temperature (°C) = l/b (log10 time/mj_n * -a)
2° C, acclimation at the maximum weekly average temperature
for summer growth, and data in Appendix II-C of Water
Quality Criteria, published by National Academy of Sciences.
c - Based on data for larvae (ERL-Duluth, 1976).
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28G
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
f lounderfPseudpj^leuronectes §J2££i£anus, in Narragansett Bay,
Rhode Island, revealed that a 78 percent decrease in annual catch
correlated closely with a Q.5°C 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 yellowtail flounder,
Limanda ferruginea, with a 1°C long-term elevation in southern
New England waters.
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, 288
TABLE 12.
Summary of Reported values for
Maximum Weekly Average Temperature for Spawning and Short-Term
Maxima for Embryo Survival During the Spawning Season
(Centigrade and Fahrenheit)
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^
5
17
25
9
21
27
10
24
21
3
21
11
9
10
17
17
10
18
18
17
18
10
12
(41)
(63)
(77)
(48)
(70)
(81)
(50)
(75)
(70)
(37)
(70)
(52)
(48)
(50)
(63)
(63)
(50)
(64)
(64)
(63)
(64)
(50)
(54)
Embryo
Survival^
7
27
34
13
33
29
13
28
26
8
27
19
13
21
21
13
24
34
26
23
20
20
(45)
(81)c
(93)
(55)
(91)
(84)
(55)
(82)c
(79)
(46)
(81)
(66)
(55)
(70)
(70)
(55)
(75)
(93)
(79)
(73)
(68)
(68)
- the optimum or mean of the range of spawning temperatures
reported for the species (ERL-Duluth, 1976).
- the upper temperature for successful incubation and
hatching reported for the species (ERL-Duluth, 1976).
- upper temperature for spawning.
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289
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 2°C. 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
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
eduljLs, sea urchin, Stroncjy_l^ocentrotus 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 (Neill 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 3°C (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, 1373). A weight loss of
approximately 10 percent for each 1° C rise in water temperature
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291
has been observed in fish when food is absent. (Phillips et al.,
1960) There nay 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 ofan-
adromous 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 Malteeos, 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).
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
-------
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 (01 la 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
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233
2j 3j7,8-TETRACHUORODIBENZO-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
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 organisms, the
ambient water concentration should be zero. This criterion is
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294
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. 5i31» February 15, 1984)
SEE APPENDIX B FOR METHODOLOGY
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TETRACHLOROETHYLSNE
CRITERIA:
Aquatic Life
The available data for tetrachloroethy1ene 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 10 , 10" , and 10 .
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
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286
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
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207
THALLIUM
CRITERIA;
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
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
ug/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.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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TOLUENE
CRITERIA;
Aquatic Life
The available data for toluene indicate that 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.
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
mg/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
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 0.3 mg/kg/day for toluene.
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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 tine. Therefore, the levels
which may result in incremental increase of cancer risk over the
lifetime are estimated at 10 , 10 , and 10 . 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 0.07 ng/L,
respectively.
(45 F.R. 79318, November 28, 1980)
SEE APPENDIX B FOR METHODOLOGY
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300
TRICHLOROETHYLENE
CRITERIA;
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.
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 1Q~5, 10~6, and 10~7.
The corresponding recommended criteria are 27 ug/L, 2.7 ug/L, and
0.27 ug/L, respectively. If these estimates are made for
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301
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
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ao;
VINYL CHLORIDE
CRITERIA:
Aguatic 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 , 10 , and
io~7. 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
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303
ZINC
CRITERIA;
Aquatic Life
For total recoverable zinc the criterion to protect freshwater
atjuatic 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 numerical value given by
e (0.83[In(hardness)]+1.95) at any time. For example, at
hardnesses of 50, 100, and 200 mg/L as Caco 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 170 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 ng/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
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30-1
APPENDIX A
-------
305
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 anight 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 for adverse effects that might
-------
30G
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 bioaccumulation
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 (publication,
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
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scr
•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
eaulsifiable 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 shc-uld 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
substantial concentrations of the test
material or other contaminants.
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308
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.
II. 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.
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
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309
IV) with at least one species of freshwater
animal in at least eight different families
such that all of the following are included:
a. the family Salmonidae in the class
Osteichthyes
b. a second family in the class
Osteichthyes, preferably a commercially or
recreationally important warmwater species
(e.g., bluegill, channel catfish, etc.)
c. a third family in the phylum Chordata (may
be in the class Osteichthyes or may be an
amphibian, etc.)
d. a planktonic crustacean (e.g., cladoceran,
copepod, etc.)
e. a benthic crustacean (e.g., ostracod,
isopod, amphipod, crayfish, etc.)
f. an insect (e.g., mayfly, dragonfly,
damselfly, stonefly, caddisfly, mosquito,
midge, etc.)
g. a family in a phylum other than Arthropoda
or Chordata (e.g., Rotifera, Annelida,
Mollusca, etc.)
h. a family in any order of insect or any
phylum not already represented.
2. Acute-chronic ratios (see Section VI) with
species of aquatic animals in at least three
different families provided that of the three
species:
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310
* 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 determined 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
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311
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
criterion can usually be derived, except in special
cases. For example, derivation of a criterion
-------
might not be possible if the available acute-
chronic 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.
III. Final Acute Value
A, Appropriate measures of the acute (short-term)
toxicity of the material to a variety of species 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
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313
Mean Acute Value of a commercially or
recreationally 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
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.
1. 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;
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314
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 IC50 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.
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 96-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.
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315
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 IC50
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 IC50.
4. Tests with single-celled organisms are not
considered acute tests, even if the duration
was 96 hours or less.
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
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316
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 or more other life stages of the
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,
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317
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.
NOTE: The geometric mean of N numbers is the N*"*1 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.
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318
NOTEs 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 for 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 "1" for the lowest to "H" for the highest.
If two or more Genus Mean Acute Values are
identical, arbitrarily assign them successive
ranks.
M. Calculate the cumulative probability, P, for each
Genus Mean Acute Value as R/(N+1).
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319
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).
0. Using the selected Genus Mean Acute Values and Pa,
calculate:
S2= E(ln GMAV)2)- ((Eln GMAV))2/4)
(P) - ((E /Ap))2/4J
L = (E(ln GMAV) - S(E(/^
A = S(/A0.05) +L
FAV = eA
(See [11] for development of the calculation
procedure and Appendix 2 for example calculation
and computer program.)
NOTE: Natural logarithms (logarithms to base e, 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
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3SO
Value.
Q. Go to Section VI.
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321
IV. Final Acute Equation
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
understanding 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 for 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
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322
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, talcing 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
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O*}
3O
iO
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 individually
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.
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324
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.
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 1,1
in the center of the graph.
H. For each species calculate the geometric mean, W,
of the acute toxicity values and the geometric
mean, X, 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,
Z, of the water quality characteristic using the
equation:
Y = in W - V{ln X - In Z).
J. For each species calculate the SMAV at Z using
the equation: SHAV = e^.
NOTE: Alternatively, the Species Mean Acute Values at Z
can be obtained by skipping step H using the
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325
equations in steps I and J to adjust each acute
value individually to Z, and then calculating the
geometric mean of the adjusted values for 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 IV.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
2, 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 - e(V[ln(water quality
characteristic)] 4- in A - V[ln Z]), where V =
pooled acute slope and A - Final Acute Value at Z.
Because V, A, and Z 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
sane manner as the Final Acute Value or by
dividing the Final Acute Value by the Final Acute-
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326
Chronic Ratio. In some cases it may not be
possible to calculate a Final Chronic Value.
NOTE: 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 acpatic 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.
B. Chronic values should be based on results of flow-
through (except renewal is acceptable for
daphnids) chronic tests in which the
concentrations of test material in the test
solutions were properly measured at appropriate
times during the test.
C. Results of chronic tests in which survival,
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a;
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
natter, etc., do not affect toxicity.
1. 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
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328
with young less than 24 hours old and last for
not lees than 21 days. Tests with mysids
should begin with young less than 24 hours old
and continue until 7 days past the median time
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 life-
cycle 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,
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329
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 after
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
life-cycle tests.
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330
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
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331
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.I or
III.C.I, 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 IV.J-O.
Then go to Section VI.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
results of all acceptable flow-through (except
static is acceptable for daphnids) acute tests in
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332
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
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33,:
whose Species Mean Acute Values are close to
the Final Acute Value.
2, If no major trend is apparent and the acute-
chronic ratios for a number of species are
within a factor of 10, the Final Acute-
Chronic Ratio should be calculated as the
geometric mean of all the Species Mean Acute-
Chronic 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
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 life-
cycle 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 XI.BJ
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334
4. If the most appropriate Species Mean Acute-
Chronic 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 XI.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 VILA.
M. If the Species Mean Chronic Value of a
commercially or recreationally 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.
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33
VI. Final Chronic Equation
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 for
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
raean of the available Species Mean Acute-
Chronic 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 VII.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
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336
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
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33"
salinity, no transformation or a different
transformation might fit the data better, and
appropriate changes will be necessary throughout
this section. It is probably preferable, but not
necessary, to use the same transformation that was
used with the acute values in Section V.
D, Decide whether the data for each species 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, chronic 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 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
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338
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 sane 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 l.o.
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
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339
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.
H. Treat all the normalized data as if they were all
for the sane 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 §5 percent confidence
limits. If all the normalized data are actually
plotted, the line of best fit will go through the
point 1,1 in the center of the graph.
I. For each species calculate the geometric mean, M,
of the toxicity values and the geometric mean, P,
of the values of the water quality characteristic.
(These were calculated in steps I and P.)
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J. 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).
NOTE: Although it is not necessary, it will usually be
best to use the sane value of the water quality
characteristic here as was used in section V.I.
K. For each species calculate a Species Mean Chronic
Value at Z using the equation: SKCV - e^.
NOTE: 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 Z, 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.
L. Obtain the Final Chronic Value at Z by using the
procedure described in Section IV.J-O.
M. If the Species Mean Chronic Value at Z of a
commercially or recreationally important species
is lower than the calculated Final Chronic Value
at 2, then that Species Mean Chronic Value should
be used as the Final Chronic Value at Z instead of
the calculated Final Chronic Value.
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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 Z. Because L, S and Z are known, the Final
Chronic Value can be calculated for any selected
value of the water quality characteristic.
VII. 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
1DTA, that might affect the toxicity of the metal.
Concentrations of EDTA above about 200 ug/L should
probably be considered excessive.
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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 bioconcentration 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.
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B. Biocon centration Factors (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 almost has to be measured in a laboratory
test. Some uptake during the bioconcentration test
might not be directly from water 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 real-
world 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
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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
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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 BCP
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 BCP, multiply the dry weight
BCF by 0.1 for plankton and by 0.2 for
individual species of fishes and invertebrates
[17].
5. If more than one acceptable BCP 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
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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
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[18-20], extrapolations can be made from tested
tissues or species to untested tissues or species
on 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 for 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 SCFs 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
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appropriate percent lipids value is 100.
b. For an FDA action level for fish, the
appropriate percent lipida value is 11
for freshwater criteria and 10 for
saltwater criteria because FDA action
levels are applied on a species-by-
species 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 froa a chronic feeding study or a
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.
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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
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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 iv.D and VI.D), from chronic tests in which
the concentrations were not measured (see VLB), from
tests with previously exposed organisms (see II.F),
and from tests on formulated mixtures or emulsifiable
concentrates (see II.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
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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)
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 "freshwater" or "saltwater"
(2) - insert name of material
(3) = insert the Criterion Continuous
concentration
(4) = insert the Criterion Maximum
Concentration,
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XII. Final Review
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 10?
4. Is the range of Species Mean Acute Values for
any genus greater than a factor of 10?
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
Values questionable?
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 recreational ly
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?
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loo
10. Are chronic values available for acutely
sensitive species?
11. Is the range of acute-chronic ratios greater
than a factor of 10?
12. Is the Final Chronic Value reasonable in
comparison with the available acute and
chronic data?
13. Is the measured or predicted chronic value for
any commercially or recreationally important
species below the Final Chronic Value?
14. Are any of the other data important?
15. Do any data look like they night be outliers?
16. 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.
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APPENDIX B
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SUMMARY 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 aninals 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.
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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 FDA action level or can be the result of a chronic wildlife
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.
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 solids,
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
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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 Hunan Health
Interpretation of the Human Health Criteria
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.
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Health assessments typically contain discussions of four
elements: exposure, pharmacokinetics, toxic effects, and
criterion f omul at ion.
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 fron 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,
nonaguatic 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 aguatic 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
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of a process to adapt/modify human health-based criteria to local
conditions, sonewhat 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.
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 nuraber.
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
protective criteria were published.
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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
pollutants 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, EPA's policy is that there is
no scientific basis for estimating "safe" levels for carcinogens.
The criteria for carcinogens, therefore, state that the
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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 1Q_7
to 10_5 (one additional case of cancer in populations ranging
from 10 million to 100,000, respectively). Other concentrations
representing different risk levels nay 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.
Suaaary of the Human Healtli 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
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
individual; the average daily consumption of freshwater and
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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 toKic manifestations. The following 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
originally used in the proposed criteria documents. The linear
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
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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.
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
for 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-effect-level data from animal
studies although human data are used wherever available. The ADI
is calculated using safety factors to account for uncertainties
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inherent in extrapolation from animal to raan. 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 Stokinger-
Woodward 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).
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APPENDIX C
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THE PHILOSOPHY OF THE 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 jun sjLt^u 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 of water, but also to protect life that is dependent
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.
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What is essential and significant life in water? Do Daphnia
or stonefly nymphs qualify as such life? Why does l/100th of a
concentration that is lethal to 5O percent of the test organisms
(LC50) constitute a criterion in some instances, whereas 1/2 or
I/10th 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.
Although experimentation relating to the effects of
particular substances under controlled conditions began in the
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early 1900'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 links 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 nay 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
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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" 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 for 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:
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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 egual 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. If eggs fail to develop 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 nay 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 food chain. Thus, fish
may mobilize lethal toxicants from their fatty tissues during
periods of physiological stress. Igg 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.
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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 subl.ethally 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 floe
or gel may be an irritant or clog fish gills to effect
asphyxiation? iron at low concentrations can be a trace nutrient
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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
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 long-
term-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
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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 quality 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 quality, 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
quality criteria involve drinking water standards, the ocean
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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
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
that peripheral portion receiving the diffuse inflow.
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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
waoi-or coldwater associated biota.
Criteria are presented for 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.
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