QUALITY
CRITERIA
FOR WATER
U.S. ENVIRONMENTAL PROTECTION AGENCY
Washington, D.C. 20460
Environmental Protection Agency
Region V, Library
230 South Dearborn Street
Chicago, Illinois 60604
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CONTENTS
Page
Foreword "*
Preface - - - ***•
Preparation of this Volume **
The Philosophy of Quality Criteria 1
Aesthetics 10
Alkalinity H
Ammonia 16
v Arsenic 25
v- Barium 36
Beryllium 39
Boron 47
v Cadmium - - - 50
Chlorine 61
v Chromium 69
Conform Bacteria 79
Color 101
Copper 107
Cyanide 128
Gases, Total Dissolved 139
Hardness 147
Iron 152
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CONTENTS CONT'D
Page
Lead ....................... - .................................. 159
Manganese -----------------------------------------------------
Mercury — ..... - ............ - ........ ---- ........... ---- ....... 183
Mixing Zones ........................ - .......... ---------------- 193
Nickel ...... - .................... - ..... - ....... — .......... — 196
Nitrates; Nitrites ................. ------ ..................... — 201
Oil and Grease - ................. ----- .............. - ......... — 21°
Oxygen, Dissolved ---------------------------------------------- 224
Pesticides:
Aldrin-Dieldrin ----------- ....... ~ ...................... 23°
Chlordane ................ — ..... - ...... - ........... ----- 24°
Chlorophenoxy Herbicides --------------------------------- 250
DDT ............... - ............................... - ...... 254
Demeton ........... ---- ................................... 26°
Endosulfan ---- ....... - ........... - .................... — - 265
Endrin ........................... ------ ...... — ....... -— 27°
Guthion .......................... - .............. ---------- 276
Heptachlor ....................... - ..................... — 284
Lindane ............................................ — ..... 291
Malathion ................. - ...... - ........ ------ ..... ----- 29&
Methoxychlor ........ - ................................ ----- 306
Mirex ............... — ....... ----- .................... — 312
Parathion .......................................... ------ 32°
Toxaphene ------------------------------------------------
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CONTENTS CONT'D
Page
pH ................................................ 337
Phenol- .......................................... 347
Phosphorus ---------------------------------------- 352
Phthalate Esters ..................................
Polychlorinated Blphenyls .........................
Selenium ------------------------------------------
Silver ........................ - ................... 388
Solids (Dissolved) & Salinity .....................
Solids (Suspended) & Turbidity ....................
Sulfides, Hydrogen Sulfide ........................
Tainting Substances -------------------- - -----------
Temperature ---------------------------------------
Zinc ..............................................
GLOSSARY ................ - .......... - .............. f98
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TAHEJES
Page
1 Maxiiiura alkalinity in waters used as a source
of supply prior to treatment «4
2. Maximum color of surface waters that have been
used as a source for industrial water supplies 104
3. The acute toxicity of copper to several species
of fish in water of various water qualities 112
4. Maximum hardness levels accepted by industry
as a raw water .source 150
5. The acute toxicity of lead to several species
of fish in water of various water qualities 166
6. Summary of lethal toxicities of various
petroleum products to aquatic organisms 212
7. Summary of some sublethal effects of petroleum
products on marine life 213
8 . Derivation of approval limits (AL) for
chlorophenoxy herbicides 252
9 . Dissolved solids hazard for irrigation water (mg/1).. 399
10. Total dissolved solids concentrations of surface
waters that have been used as sources for
industrial water supplies 401
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) 481
12. Summary values for maximum weekly average
temperature for spawning and short-term
maxima for embryo survival during the spawning
season (centigrade and fahrenheit) 434
13. Selected Thermal Requirenents and Limiting Tenperature
Data Gold Twperate Zone, Atlantic Coast: South of
long Island, N.Y. to Cape Hatteras, N.C. .-.
14-• The acute toxicity (24-, 48-, 96-hour TLso values)
of zinc to several species of fish in water of
various water qualities
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FIGURES
Page
1. Median Resistance Times To High Temperatures 429
2. Graph to determine the maximum weekly average
temperature of plumes for various ambient
temperatures, °C (°F) 432
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UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON. D.C. 20460
2 6 JUL 1376
OFFICE OF WATER AND
HAZARDOUS MATERIALS
To the Reader:
Thousands of fine scientists throughout the country have
contributed directly or indirectly to this publication on
"Quality Criteria for Water." This volume represents a stock-
taking effort on the part of this Agency to identify as precisely
as possible at this time, on a national scale, the various water
constituents that combine to form the concept of "Quality Criteria
for Water". This process of definition will continue far into
the future because research related to water quality is a never-
ending evolutionary process, and the water environment is so
complex that man's efforts to define it will never attain finite
precision.
Water quality criteria do not have direct regulatory use, but
they form the basis for judgment in several Environmental Protection
Agency and State programs that are associated with water quality
considerations. The criteria presented in this publication should not
be used as absolute values for water quality. As it is stated in the
chapter on "The Philosophy of Quality Criteria" there is variability
in the natural quality of water and certain organisms become adapted
to that quality, which may be considered extreme in other areas.
These criteria represent scientific judgments based upon literature
and research about the concentration-effect relationship to a particular
water quality constituent upon a particular aquatic species within the
limits of experimental investigation. They should be used with con-
sidered judgment and with an understanding of their development. The
judgment associated with their use should include the natural quality
of water under consideration, the kinds of organisms that it contains,
the association of those species to the particular species described
in this volume upon which criteria values have been placed, and the
local hydro!ogic conditions.
It must be emphasized that national criteria can never be de-
veloped to meet the individual needs of each of the Nation's water-
ways—the natural variability within the aquatic ecosystem can never
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be identified with a single numerical value. Water quality criteria
will change in the future as our knowledge and perception of the
intricacies of water improve. There is no question but that criteria
for some constituents will change within a period of only two years
based upon research now in progress. That is a mark of continuing
progressive research effort, as well as a mark of a better under-
standing by man of the environment that he inhabits.
This, then, is the challenge for the future: to expand upon our
present baseline of knowledge of the cause-effect relationships of
water constituents to aquatic life and of the antagonistic and syner-
gistic reactions among many quality constituentsJn water; and to mold
such future knowledge into realistic, environmentally protective
criteria to insure that the water resource canxfalfill society's
needs.
:kardt C. Beck
Deputy Assistant Administrator
for Water Planning and Standards
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FOREWORD
The Federal Water Pollution Control Act Amendments of 1972 require
the Administrator of the Environmental Protection Agency to publish
•criteria for water quality accurately reflecting the latest scientific
knowledge on the kind and extent of all Identifiable effects on health
and welfare which may be expected from the presence of pollutants in any
body of water, Including ground water. Proposed Water Quality Criteria
were developed and a notice of their availability was published on
October 26, 1973 (38 FR 29646). This present volume represents a
revision of the proposed water quality criteria based upon a consideration
of comments received from other Federal agencies, State agencies, special
Interest groups, and Individual scientists.
This volume, Quality Criteria for Water, addresses the effects of
those basic water constituents and pollutants that are considered most
significant In the aquatic environment 1n the context of our present
knowledge and experience. The format for criteria presentation has been
altered substantially from the proposed volume. It 1s believed that the
alphabetical arrangement of the water quality constituents and the form
1n which the Information Is arranged will be of considerable help to the
reader in using this volume. For each basic water constituent or pollu-
tant there 1s a recommended criterion, an Introduction, a rationale
supporting the recommended criterion and a 11st of the references cited
in the development of the recommendation.
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The thrust of this volume is to recommend criterta levels for a
water quality that will provide for the protection and propagation of
fish and other aquatic life and for recreation in and on the water in
accord with the 1983 goals of P.L. 92-500. Criteria also are presented
for the domestic water supply use. Generally, these water uses are the
highest achievable beneficial uses and water quality that supports these
uses will also be suitable for agricultural and industrial uses. In
those few exceptions, criteria are presented to provide a safe water
quality for agricultural use, or water quality conditions associated
with agricultural and industrial uses are discussed in the rationale
supporting.a criterion recommendation.
Guidelines to implement the consideration of criteria presented in
this volume in the development of water quality standards, and in other
water-related programs of this Agency, are being developed and will be
available for use by the States, other agencies, and interested parties
In the near future.
jssell E.
'Administrator
I/
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PREFACE
The genesis of water quality criteria in the United States began in the
early 1900's. Marshl/, in 1907, published on the effects of industrial
wastes on fish. Shelford2-/, in 1917, published effect data on fish for a
large number of gas-waste constituents. In this early publication he
reiterated that the toxicity of waste differs for different species of fish
and generally is greater for the smaller and younger fish. Powers—', working
with Shelford, experimented with the goldfish as a test animal for aquatic
toxicity studies.
A monumental early effort to describe and record the effects of various
concentrations of a great number of substances on aquatic life was that of
Ellis in 1937. Ellis^/ reviewed the existing literature for 114 substances
in a 72-page document and listed lethal concentrations found by the'various
authors. He provided a rationale for the use of standard test animals in
aquatic bioassay procedures and he used the common goldfish, Crassius auratus
and the entomostracan, Daphnia magna, as test species in which experiments
were made in constant temperature cabinets.
Early efforts to summarize knowledge concerning water quality criteria
-' M.C. MARSH, The effect of some industrial wastes on fishes. Water supply
and irrigation paper No, 192, U.S. Geol. Sur., pp. 337-348 (1907).
y V.E. SHELFORD, An experimental study of the effects of gas wastes upon
fishes, with especial reference to stream pollution. Bull. Illinois
State Lab. for Nat. History, n_:381-412 (1917).
I/ E.B. POWERS, The goldfish (Carassius carassius) as a test animal in the
study of toxicity. Illinois Biol. Mono., 4_: 127-193 (1917).
i/ M.M. ELLIS, Detection and measurement of stream pollution. Bull. U.S.
Bureau of Fisheries, 48:365-437 (19.37).
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took the form of a listing of the concentration, the test organism, the results
of the test within a time period, and the reference for a cause-effect relation-
ship for a particular water contaminant. In early bioassay efforts insufficient
attention was given to the quality of the dilution-water used for the experi-
ment and to the effects of such dilution water on the relative toxicity of the
tested contaminant. As a result, conclusions from citations of such references
were, at best, difficult to formulate and most often were left to the discretion
of the reader.
In 1952, the State of California!/ published a 512-page book on "Water
Quality Criteria"' that contained 1,369 references. This classic reference
summarized water quality criteria promulgated by State and interstate agencies,
as well as the legal application of such criteria. Eight major beneficial
uses of water were described. Three-hundred pages of the document were
devoted to cause-effect relationships for major water pollutants. The
concentration-effect levels for the pollutant in question were discussed for
each of the designed water uses.
The State of California's 1952 Water Quality Criteria edition was expanded
and tremendously enhanced into a second edition edited by Messrs. Jack E. McKee
and Harold W. Wolf and published in 1963 by the Resources Agency of California,
State Water Quality Control Boards/. This edition, which included 3,827 cited
references, was a monumental effort in bringing together under one cover the
world's literature on water quality criteria as of the date of publication.
Criteria were identified and referenced for a host of water quality
Water Quality Criteria. State Water Pollution Control Board, Sacramento,
California.
J.E. McKEE and H.W. WOLF, Water Quality Criteria, State Water Quality
Control Board, Sacramento, California, Pub. 3-A -(1963)
tV
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characteristics according to their effects on domestic water supplies,
industrial water supplies, irrigation waters, fish and other aquatic life,
shellfish culture, and swimming and other recreational uses. Specific values
were arranged in ascending order, with appropriate references, as they had been
reported damaging to fish or as not harmful to fish in the indicated time and
under the conditions of exposure. The results of such a tabulation presented a
range of values and, as would be expected by those investigating such conditions,
there was often an overlap in values between those concentrations that had been
reported by others as harmful. Such an anomaly is due to differences in investi-
gative techniques among investigators, the characteristics of the water used as
a dilutent for the toxicant, the physiological state of the test organisms,
and variations in the temperature under which the tests were conducted. Neve* •
the less, the tabulation of criteria values for each of the water quality con-
stituents has been helpful through time to predict a range within which a water
quality constituent would have a deleterious effect upon the receiving waterway.
In 1966 the Secretary of the Interior appointed a number of nationally
recognized scientists to a National Technical Advisory Committee to develop water
quality criteria for five specified uses of water: domestic water supply,
recreation, fish and wildlife, agricultural, and industrial. In 1968 the report
was published^/. This report constituted the most comprehensive documentation
to date on water quality requirements for particular and defined water uses.
The book was intended to be used as a basic reference by personnel in State
water pollution control agencies engaged in water quality studies and water
quality standards setting activities. In some respects, this volume represented
Water Quality Criteria, A Report of the National Technical Advisory
Committee to the Secretary of the Interior. U.S. Government
Printing Office, Washington, D.C. (1968).
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a marriage between the best available experimental or investigative criteria
recorded in the literature and the judgments of recognized water quality experts
with long experience in associated management practices. Its publication heralded
a change in the concepts of water quality criteria from one that listed a series
of concentration-effect levels to another that recommended concentrations that
would ensure the protection of the quality of the aquatic environment and the
continuation of the designated water use. When a specific aquatic life recom-
mendation for a particular water pollutant could not be made because of either a
lack of information or conflicting information, a recommendation was made to
substitute a designated application factor based upon data obtained from a 96-hour
bioassay using a sensitive aquatic organism and the receiving water as a diluent
for the toxicity test.
The U.S. Environmental Protection Agency contracted with the National
Academy of Sciences and the National Academy of Engineering to embellish the
concept of the 1966 National Technical Advisory Committee's Water Quality Criteria
and to develop a water quality criteria document that would include current
knowledge. The result was a 1974 publication that presented water quality
criteria as of 1972^.
The Federal Water Pollution Control Act Amendments of 1972 (P.L. 92-500)
mandated that the Environmental Protection Agency publish water quality criteria
accurately reflecting the latest scientific knowledge on the kind and extent of
all identifiable effects on health and welfare which may be expected from the
presence of pollutants in any body of water.
Water Quality Criteria, 1972. National Academy of Sciences, Natiorral
Academy of Engineering. U.S. Government Printing Office,
Washington, D.C. (1974).
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Section 304(a) of P.L. 92-500 states, "(1) The Administrator, after
consultation with appropriate Federal and State agencies and other interested
persons, shall develop and publish, within one year after October 18, 1972 (and
from time to time thereafter revise) criteria for water quality accurately
reflecting 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, esthetics,
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 rates of organic and inorganic sedimentation for
varying types of receiving waters.
"(2) The Administrator, after consultation with appropriate Federal and
State agencies and other interested persons, shall develop and publish, within
one year after October 18, 1972 (and from time to time thereafter revise)
information (A) on the factors necessary to restore and maintain the chemical,
physical, and biological integrity of all navigable waters, ground waters,
waters of the contiguous zone, and the oceans; (B) on the factors necessary for
the protection and propagation of shellfish, fish, and wildlife for classes and
categories of receiving waters and to allow recreational activities in an on
the water; and (C) on the measurement and classification of water quality;
and (D) for the purpose of Section 303 of this title, on and the identification
of pollutants suitable for maximum daily load measurement correlated with the
achievement of water quality objectives.
V//
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"(3) Such criteria and information and revisions thereof shall be issued
to the States and shall be published in the Federal Register and otherwise
made available to the public."
Section 101(a)(2) of P.L. 92-500 states, "It is the national goal that
wherever attainable, an interim goal of water quality which provides for the
protection and propagation of fish, shellfish, and wildlife, and provides for
recreation in and on the water, will be achieved by July 1, 1983."
The objectives of this volute are to respond to these
sections of the Act and thus establish water quality criteria".
The QCW will be expanded periodically in the future to include
additional constituents as data become available. While the
NAS/ftAE 1972 Water Quality Criteria considered aluminum,
antimony, bromine, cobalt, fluoride, lithium, molybdenum,
thallium, uranium and vanadium, these presently are not included
in this volume; however, they should be given consideration in
the development of State Water Quality Standards and quality
criteria may be developed for them in future volumes of the QCW.
In particular geographical areas or for specific water uses such
as the irrigation of certain crops, some of these constituents
may have harmful effects. Until such time that criteria for the
10 aforementioned constituents are developed, information
relating to their effects on the aquatic ecosystem may be found
in the NAS/NAE 1972 Water Quality Criteria.
vnt
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PREPARATION OF THIS VOLUME
A volume of this scope results from the efforts of
many dedicated people and Includes technical specialists
located'tnrbughout the Agency's operational programs and
1n Its research laboratories. The responsibility for
coordinating compilation efforts and in preparing manu-
script was assigned to the Criteria Branch of the Criteria
and Standards Division within the Office of Water Planning
and Standards, EPA.
IX
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THE PHILOSOPHY OF QUALITY CRITERIA
Water quality criteria specify concentrations of water constituents which,
if not exceeded, are expected to support an aquatic ecosystem suitable for
the higher uses of water. Such criteria are derived from scientific facts
obtained from experimental or in situ observations that depict organism
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
safety for survival and propagation at all times to all organisms within a
given ecosystem. They are intended not only to protect essential and sig-
nificant life in water, as well as 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.T.
300f, et seg^.), 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 1n water? Do Daphnia or stonefly
nymphs qualify as such life? Why does 1/100th of a concentratiort that 1s
lethal to 50 percent of the test organisms (LC50) constitute a criterion in
some Instances, whereas l/20or l/10th of some effect levels constitute 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 1s 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 early 1900's, the effects of any sub-
stance on more than a few of the vast number of aquatic organisms have not
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been investigated. Certain test animals have been selected by investigators
for intensive investigation because of their importance to man, because of
their availability to the researcher and because of their physiological re-
sponses 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 to protect associated organisms.within
the aquatic environment that have not been tested and that may be more
sensitive to the test constituent. Application factors have been used to
provide the degree of protection required. Safe levels for certain chlorinated
hydrocarbons and certain heavy metals were estimated by applying an 0.01
application factor to the 96 hour LC50 value for sensitive aquatic organisms.
Flow-through bioassays have been conducted for some test indicator organisms
nv«?r a substantial norfnH o* their li-pp historv. 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.
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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; often death is the end point.
Acute toxicity can be expressed as the lethal concentration for a stated
percentage of organisms tested, or the reciprocal, which is the tolerance
limit of a percentage of surviving organisms. Acute toxicity for aquatic
organisms generally has been expressed for 24- to 96-hour exposures.
Chronic toxicity refers to effects through an extended time period.
Chronic toxicity may be expressed in terms of an observation period equal
to the lifetime of an organism or to the time span of more than one
generation. Some chronic effects may be reversible, but most are not.
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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 re-
productive failure. Physiological stress may make a species less competitive
with others and may result in a gradual population decline or absence from
an area. The elimination of a microcrustacean that serves as a vital food
during the larval period of a fish's life could result ultimately in the
elimination of the fish from an area. The phenomenon of bioaccumulation of
certain materials may result in chronic toxicity to the ultimate consumer in
a food chain. Thus, fish may mobilize Ipthal toxicants frop their fatty
tissues during periods of physiological stress. Egg shells of predatory
birds may be weakened to a point of destruction in the nest. Bird chick embryos
way have increased mortality rates. There may be a hazard to the health of
man if aquatic organisms with toxic residues are consumed.
The fact that living systems, i.e. individuals, populations, species
and ecosystems can take up, accumulate, and bioconcentrate man-made and natural
toxicants is well documented. In aquatic systems biota are exposed directly
to pollutant toxicants through submersion in a relatively efficient
solvent (water) and are exposed indirectly through food webs and other
biological, chemical, and physical interactions. Initial toxicant levels,
if not immediately toxic and damaging, may accumulate in the biota or
sediment over time and increase to levels that are lethal or sublethally
damaging to aquatic organisms or to consumers of these organisms. Water
quality criteria reflect a knowledge of the capacity for environ-
mental accumulation, persistence, and effects of specific toxicants in
specific aquatic systems.
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Ions of toxic materials frequently cause adverse effects because they
pass through the semi permeable membranes of an organism. Molecular diffusion
through membranes may occur for some compounds such as pesticides, poly-
chlorinated blphenyls 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 but at high con-
centrations 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 environments 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
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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 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
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 character!stlcr, and synerglstic or antagonistic
effects of combinations of pollutants.
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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 make use of water quality criteria include drinking water
standards, the ocean dumping program, designation of hazardous substances,
dredge spoil criteria development, removal of in-place toxic materials,
thermal pollution, and pesticide registration.
To provide the water resource protection for which they are designed,
quality criteria should apply virtually to 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 effect 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/ft 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 aquatic plscicides 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 1t 1s
believed that such diffuse source inflows should receive consideration
e
-------
similar to that of a discrete effluent discharge and that 1n such Instances
the criteria should be applied to the principal portion of the waterway rattet than
to that peripheral portion receiving the diffuse inflow.
The format for presenting water quality criteria includes a concise
statement of the dominant criterion or criteria for a particular constituent
followed by a narrative introduction, a rationale that includes justification
for the designated criterion or criteria, and a listing of the references
cited within the rationale. An effort has been made to restrict supporting
data to those which have either 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, pro-
vision for recreational bathing or water skiing, and permitting an appropriate
factor of safety to ensure protection for essential warm or cold water
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 of the 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.
-------
AESTHETIC QUALITIES
CRITERIA:
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
matter 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 1n the
public interest.
Aesthetic qualities provide the general rules to protect water against
environmental insults; they provide minimal freedom requirements from pollution;
they are essential properties to protect the nation's waterways.
10
-------
CRITERION;
20 mg/1 or more as CaCO, for freshwater aquatic life except
where natural concentrations are less.
INTRODUCTION:
Alkalinity is the sum total of components in the water that tend
to elevate the pH of the water above a value of about 4.5 It is
measured by titration with standardized acid to a pH value of about
4.5 and it is expressed commonly as milligrams per liter of calcium
carbonate. Alkalinity, therefore, is a measure of the buffering capacity
of the water, and since pH has a direct effect on organisms as well
as an indirect effect on the toxicity of certain other pollutants in the
water, the buffering capacity is important to water quality. Examples
of commonly occurring materials in natural waters that increase the
alkalinity are carbonates, bicarbonates, phosphates and hydroxides.
RATIONALE:
The alkalinity of water used for municipal water supplies is important
because it affects the amounts of chemicals that need to be added to accomplish
coagulation, 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 alkalirlity do not have to be supplemented with arti-
ficial ly'added materials to increase the alkalinity. Alkalinity resulting from
naturally occurring materials such as carbonate and bicarbonate is not considered
a health hazard in drinkinq water supplies, £er se_. and naturally occurring
-------
maximum levels up to approximately 400 mg/1 as calcium carbonate are not
considered a problem to human health (NAS, 1974).
Alkalinity is important for fish and other aquatic life in fresh-
water 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 bicarbonate will complex some
toxic heavy metals and reduce their toxicity markedly. For these reasons, the
National Technical Advisory Committee (NTAC, 1968) recommended a minimum
alkalinity of 20 mg/1 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 percent reduction avoids
the problem of establishina standards on waters where natural alkalinity is
at or below 20 mg/1. 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 h-is been
noted as a correlation that productive waterfowl habitats are above 25 mg/1
with higher alkalinities resulting in better waterfowl habitats (NTAC, 1968).
Excessive alkalinity can cause problems for swimmers by altering the pH of
the lacrimal fluid around the eye, causing irritation.
For industrial water suppli ;s 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 pl^nt- ^-*"age. 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 water and make the iron unavailable to plants. Such deficiencies induce
chlorosis and further plant damage. Usually alkalinity must exceed 600 mg/1
before such affects are noticed, however.
/$
-------
TABLE I *
MAXIMUM ALKALINITY IN WATERS USED AS A SOURCE
OF SUPPLY PRIOR TO TREATMENT
Alkalinity
Industry mg/1 as
Steam Generation Boi 1 er Makeup 350
Steam Generati on Cool i ng 500
Textile Mill Products 50-200
Paper and Allied Products 75-150
Chemi cal and Al 1 i ed Products 500
Petroleum Refining 500
Primary Metals Industries 200
Food CAnning Industries 300
Bottl ed and Canned Soft Dri nks 85
* NAS, 1974
-------
REFERENCES CITED
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972, U.S. Government Printing Office,
Washington, D. C.
National Technical Advisory Committee to the Secretary of the Interior,
1968. Water quality criteria, U.S. Government Printing Office,
Washington, D. C.
-------
AMMONIA
CRITERION:
mg/1 (as Un-Ionized Ammonia) for freshwater aquatic life.
Concentrations.of Total Ammonia (NH3 + NH4 ) Which
Contain an Un-Ionized Ammonia Concentration of 0.020 mg/s,
Temper-
ature
CC)
5
10
15
20
25
30
pH Value
•
6.0
160.
no.
73.
50.
35.
25.
6.5
51.
34.
23.
16.
11.
7.9
7.0
16.
11.
7.3
5.1
3.5
2.5
7.5
5.1
3.4
2.3
1.6
1.1
0.81
8.0
1.6
1.1
0.75
0.52
0.37
0.27
8.5
0.53
0.36
0.25
0.18
0.13
0.099
9.0
0.18
0.13
0.093
0.070
0.055
0.045
9.5
0.071
0.054
0.043
0.036
0.031
0.028
10.0
0.036
0.031
0.027
0.025
0.024
0.022
"[Abstracted frow Thurston et a1_. (1974)]
INTRODUCTION:
Ammonia is a pungent, colorless, gaseous, alkaline compound of nitrogen
and hydrogen that is highly soluble in water. It is a biologically active
compound present in most waters as a normal biological degradation product
of nitrogenous organic matter. It may also reach ground and surface waters
through discharge of industrial wastes containing ammonia as a byproduct
or wastes from industrial processes using "ammonia water".
When ammonia dissolves in water, some of the ammonia reacts with the
water to form ammonium ions. A chemical equilibrium is established which
-------
contains un-ionized ammonia (NH3), ionized ammonia (NH, ), and hydroxide
ions (OH"). The equilibrium for these chemical species can be expressed
in simplified form by the following equation:
NH3 + H20 ^ NH3-H20 ^ NH4+ + OH"
In the above equation NH, represents ammonia gas combining with water.
The term NH-'H^O represents the un-ionized ammonia molecule which is
loosely attached to water molecules. Dissolved un-ionized
ammonia will be represented for convenience as NH_. The ionized form
of ammonia will be represented as NH4 . The term total ammonia will refer
to the sum of these (NH3 + NH4+).
The toxicity of aqueous solutions of ammonia is attributed to the NH3
species. Because of the equilibrium relationship among NH3, NH4 , and
OH", the toxicity of ammonia is very much dependent upon pH as well as
the concentration of total ammonia. Other factors also affect the con-
centration of NH3 in water solutions, the most important of which are
temperature and ionic strength. The concentration of NH3 increases with
increasing temperature, and decreases with increasing ionic strength.
In aqueous ammonia solutions of dilute saline concentrations, the NH3
concentration decreases with increasing salinity.
A table of percent NH3 for aqueous ammonia solutions of zero salinity
at different values of pH and temperature is given below. This percentage
can be used to determine the amount of total ammonia which is in the most
toxic (NH3) form.
-------
Percent Un-ionized Ammonia in Aqueous Ammonia Solutions
Temper-
ature
(°C)
5
10
15
20
25
30
pH Value
6.0
0.013
0.019
0.027
0.040
0.057
0.08C
6.5
0.040
0.059
0.087
0.13
0.18
0.25
7.0
0.12
0.19
0.27
0.40
0.57
0.80
7.5
0.39
0.59
0.86
1.2
1.8
2.5
8.0
1.2
1.8
2.7
3.8
5.4
7.5
8.5
3.8
5.6
8.0
11.
15.
20.
9.0
11.
16.
21.
28.
36.
45.
* 9.5
28.
37.
46.
56.
64.
72.
10.0
56.
65.
73.
80.
85.
89.
[Thurston, et al_. (1974)]
RATIONALE;
It has been known since early In this century that ammonia 1s toxic
to fishes and that the toxicity varies with the pH of the water. Chipman
(1934) demonstrated that undissodated ammonia (NHa) was the chemical species
toxic to goldfish, amphipods, and cladocerans. He concluded from his
studies that the toxicity of ammonium salts was pH-dependent and was directly
related to the concentration of undissociated ammonia. Chipman's work was
confirmed by Wuhrmann, e£ a].. (1947) who concluded that the NH3 fraction
was toxic to fish and that the NH^*" fraction had little or no toxicity.
Further studies by Wuhrmann and Woker (1948) and Downing and Merkens (1955)
agreed with these earlier findings. Tabata (1962), however, has attributed
some degree of toxicity to fishes and invertebrates to the NH^ species (less
than l/50th that of NH3).
-------
In most natural waters, the pH range is such that the NH. fraction
of ammonia predominates; however, in highly alkaline waters, the NFU frac-
tion can reach toxic levels. Many laboratory experiments of relatively
short duration have demonstrated that the lethal concentrations for a
variety of fish species are in the range of 0.2 to 2.0 mg/1 NHo with trout
being the most sensitive and carp the most resistant. Although coarse
^fish such as carp survive longer in toxic solutions than do salmonids, the
difference in sensitivity between fish species to prolonged exposure is
probably less than the tenfold range given as the lethal range. The lowest
lethal concentration reported for salmonids is 0.2 mg/1 NH3 for rainbow trout
fry (Li-ebmann, 1960). The toxic concentration for Atlantic salmon smol.ts
(Herbert and Shurben, 1965) and for rainbow trout (Ball, 1967) was found
to be only slightly higher. Although the concentration of NFL below 0.2
mg/1 may not kill a significant proportion of a fish population, such con-
centration may still exert an adverse physiological or histopathological
effect (Flis, 1968; Lloyd and Orr, 1969; Smith and Piper, 1974). Fromm
(1970) found that at total ammonia (NHL + NH, ) concentrations of 3 mg/1
ammonia nitrogen, the trout became hyperexcitable; at 5 mg/1, ammonia
excretion by rainbow trout was inhibited; and at 8 mg/1 (approximately
1 mg/1 NHg), 50 percent died within 24 hours. Burrows (1964) found progressive
gill hyperplasia in fingerling Chinook salmon during a six-week exposure
to a total ammonia concentration (expressed as NH4) of 0.3 mg/1 (0.002
mg/1 NH3), which was the lowest concentration applied. Reichenbach-Klinke
(•1967) also noted gill hyperplasia, as well as^pathological effects on
19
-------
the liver and blood of various species at a concentration of 0.27 mg/1
NH3. Flis (1968) noted that exposure of carp to sublethal NH3 concentra-
tions resulted in extensive necrotic changes and tissue disintegration in
various organs.
Herbert and Shurben (1965) reported that the resistance of
yearling rainbow trout to ammonia increased with salinity (i.e. dilution
with about 30 percent sea water) but above that level resistance appeared
to decrease. Katz and Pierro (1967) subjected fingerling coho salmon
to an ammonia waste at salinity levels of 20, 25, and 29 parts per
thousand (i.e. dilution with about 57-83 percent sea water) and also
found that toxicity increased with increased salinity, In saline waters
the NH4+/NH3 ratio must be adjusted by consideration of the activity of the
charged species and total ionic strength of the-solution. In dilute saline
waters this ratio will change to favor NH. , and thereby reduce the concen-
tration of the toxic NH3 species. At higher salinity levels the reported
toxic effects of ammonia to fish must therefore be attributed to some
mechanism other than changes in the NH4+/NH3 ratio. Data on the effect of
ammonia on marine species are limited and the information on anadromous
species generally has been reported in conjunction with studies on fresh-
water species.
Although the NH3 fraction of total ammonia increases with temperature,
r
the toxic effect of NH3 vs. temperature is not clear. Burrows (1964) has
reported that the recovery rate from hyperplasia in gill tissues of chinook
salmon exposed first to ammonia at sublethal levels and then to fresh water
was less at 6t than at 14°C. In this experiment, comparison was made bet.ween\
two different age classes of salmon.
-------
Levels of un-ionized ammonia in the range of 0.20 to 2 mg/1 have been
shown to be toxic to some species of freshwater aquatic life. To provide
safety for those life forms not examined, 1/1Oth of the lower value of
this toxic effect range results in a criterion of 0.020 mg/1 of un-ionized
ammonia. This criterion is slightly lower than that recommended for European
inland fisheries (EIFAC, 1970) for temperatures above 5° C and pH values
below 8.5. Measurement of values of total ammonia for calculation of
values in the range of 0.020 mg/1 NH3 is well within current analytical
capability.
-------
REFERENCES CITEDi
Ball, I.R., 1967. The relative susceptibilities of some species of fresh-
water fish to poisons - I. Ammonia. Water Res. 1: 767.
Burrows, R.E., 1964. Effects of accumulated excretory products on hatchery-
reared salmonids. Bureau of Sport Fisheries and Wildlife Research Report
66. G.P.O., Washington, D.C.
Chipman, W.A., Jr., 1934. The role of pH in determining the toxicity
of ammonium compounds. Ph.D. Thesis, University of Missouri, Columbia,
Missouri.
Downing, K.M. and O.C. Merkens, 1955. The influence of dissolved oxygen
concentration on the toxicity of un-ionized ammonia to rainbow trout
(Salmo gairdnerii Richardson). Ann. appl. Biol., 43: 243.
European Inland Fisheries Advisory Commission, 1970. Water quality criteria
for European freshwater fish. Report on ammonia and inland fisheries.
EIFAC Technical Paper No. 11, 12 p; Water Res» 7: 1011 (1973).
Flis, J., 1968. Anatomicohistopathological changes induced in carp (Cyprinus
carpio L.) by ammonia water. Part I. Effects of toxic concentrations.
Acta Hydrobiol., 10: 205. Part II. Effects of subtoxic concentrations.
Ibid., 10: 225.
Fromm, P.0», 1970. Toxic action of water soluble pollutants on freshwater
fish. EPA Water Pollution Control Research Series No. 18050DST. G.P ,,
Washington, D.C.
-------
Herbert, D.W.M. and D.S. Shurben, 1965. The susceptibility of salmonid
fish to poisons under estuarine conditions - II. Ammonium chloride.
Int. J. Air Wat. Pollut., 9: 89.
Katz, M. and R.A. Pierro, 1967. Estimates of the acute toxicity of ammonia-
urea plant wastes to coho salmon, Qncorhynchus kisutch. Final Report,
Fisheries Research Institute, University of Washington, Seattle: 15 p.
Liebmann, H.,.1960. Handbuch der Frischwasser- und Abwasserbiologie - II
Miinchen.
Lloyd, R. and L.D. Orr, 1969. The diuretic response by rainbow trout to
suWethal concentrations of ammonia. Wat. Res., 3: 335.
Reichenbach-Klinke, H.-H., 1967. Untersuchungen iiber die Einwirkung des
Ammoniakgehaltb auf den Fischorganismus. Arch. Fischereiwiss., 17: 122.
smith, C.E. and R.G. Piper, 1975. Lesions associated with chronic exposure
to ammonia. IN: The Pathology of Fishes, W.E. Ribelin and G.
Mlgaki (Eds.), University of Wisconsin Press, Madison, pp 497-514.
Tabata, K., 1962. Toxicity of ammonia to aquatic animals with reference
to the effect of pH and carbon dioxide. (In Japanese with English summary),
Bull. Tokai Reg. F1sh. Res. Lab., 34: 67.
-------
Thurston, R.V., e_t al_., 1974. Aqueous ammonia equilibrium calculations.
Fisheries Bioassay Laboratory Technical Report No. 74-1, Montana State
University, Bozeman, 18 p.
Wuhrmann, K., et^al_., 1947. liber die fischereibiologische Bedeutung des
Ammonium- und Ammoniakgehaltes fliessender Gewasser. Vjschr. naturf,
Ges. Zurich, 92: 198.
Wuhrmann, K. and H. Woker, 1948. Beitrage zur Toxikologie der Fische.
II. Experimentelle Untersuchungen u'ber die Ammoniak- und Blausaurevergiftung.
Schweiz. Z. Hydrol., 11: 210.
-------
ARSENIC
CRITERIA:
50 ug/1 for domestic water supplies (health);
100 ug/1 for irrigation of crops.
INTRODUCTION:
Arsenic is a shiny, gray, brittle element possessing both metallic
and non-metallic properties. Compounds of arsenic are ubiquitous in
nature, insoluble in water and occur mostly as arsenides and arseno-
pyrites. Samplings from 130 water stations in the United States have
shown arsenic concentrations of 5 to 336 ug/1 with a mean level of 64
ug/1 (Kopp, 1969). Arsenic normally is present in sea water at concen-
trations of 2 to 3 ug/1.
Arsenic exists in the trivalent and pentavalent states and its
compounds may be either organic or inorganic. Trivalent inorganic
arsenicals are more toxic than the pentavalent forms both to mammals
and aquatic species. Though most forms of arsenic are toxic to humans,
arsenicals have been used in the medical treatment of spirochaetal
infections, blood dyscrasias and dermatitis (Merck Index, 1968). Arsenic
and arsenicals have many diversified industrial uses such as hardening of
copper and lead alloys, pigmentation in paints and fireworks, and the
manufacture of glass, cloth and electrical semiconductors. Arsenicals
are used in the formulation of herbicides for forest management and
agriculture.
25"
-------
RATIONALE:
Arsenic concentrations in most community drinking water supplies
in the United States range from a trace to approximately 0.1 mg/1
(McCabe, et al_., 1970).
Inorganic arsenic is absorbed readily from the gastrointestinal
tract, the lungs, and to a lesser extent from the skin, and becomes
distributed throughout the body tissues and fluids (Sollman, 1957).
It is excreted via urine, feces, sweat, and the epithelium of the skin
(Ducoff, et aJL, 1948; Musil and Dejmal, 1957).
After cessation of continuous
exposure, arsenic excretion may continue for as long as 70 days (DuBois
and Ceiling, 1959).
Since the early nineteenth century arsenicals have been suspected
of being carcinogenic (Paris, 1820; Sommers and McManus, 1953; Boutwell,
1963; Hueper and Payne, 1963).
According to Frost (1967), the most toxic arsenicals are well
tolerated at concentrations of 10 to 20 ppm arsenic in the diet. The
least toxic arsenicals can be fed without injury at levels which
contribute up to at least 1000 ppm arsenic in the diet. He concluded
that arsenicals appear remarkably free of carcinogenic properties.
In man, subacute and chronic arsenic poisoning may be insidious
and pernicious. The symptoms of mild chronic poisoning are fatigue
-------
and loss of energy. In more severe intoxication the following symptoms
may be observed: gastrointestinal catarrh, kidney degeneration, tendency
to edema, polyneuritis, liver cirrhosis, bone marrow injury, exfoliate
dermatitis and altered skin pigmentation (DiPalma, 1965; Goodman and
Gilman, 1965). No true tolerance of arsenic has ever been demonstrated
(Dubois and Ceiling, 1959). During chronic exposure, trivalent arsenic accumulates
mainly in bone, muscle, and skin and to a lesser degree in the liver and kidneys
(Smith, 1967).
Reports from epidemiological studies in Taiwan indicate that
0.3 mg/1 arsenic in drinking water resulted in increased incidences of
hyperkeratosis and skin cancer with increased consumption of water
(Chen and Wu, 1962; Tseng, e_t al_., 1968; Yeh, et a]_., 1968). A similar
situation has been reported in Argentina (Trelles, e_t al_., 1970).
Dermatological manifestations of arsenicism were noted in children in
Antofagasta, Chile who used a water supply containing an arsenic
concentration of 0.8 mg/1. A new water supply was provided and preliminary
data showed that the quantity of arsenic in hair decreased (Borgono
and Grieber, 1972).
Arsenicism affecting two members of a family whose well water
concentration varied from 0.5 to 2.75 mg/1 arsenic over a period of
several months was reported in Nevada (Craun and McCabe, 1971). A
study in California indicated that a greater proportion of the
population had elevated concentrations of arsenic in hair when their
drinking water had more than 0.12 mg/1 arsenic than when concentrations
were lower, but illness was not noted (Goldsmith, ejt aK, 1972). In
none of the cited incidents of apparent correlation of arsenic in drinking
water with increased incidence of hyperkeratosis and skin cancer has there
been any confirmed evidence that arsenic was the etiologic agent in
the production of carcinomas.
-------
It is estimated that the total intake of arsenic from food
averages 900 ug/day (Schroeder and Balassa, 1966). At the concentration
of 50 ug/1 recommended for drinking water supply and an average intake
of 2 liters ot water per dav. the intake from water could reach 100
ug/day or approximately 10 percent of the total ingested arsenic.
Although arsenic is concentrated in aquatic organisms, it is
evidently not progressively concentrated along a food chain. In addition,
arsenic consumed as an organically-bound species in flesh appears to
have low toxicity (Ferguson and Gavis, 1972). Surber and Meehan (1931)
found that fish-food organisms generally can withstand concentrations of
*
approximately 1.73 mg/1 of arsenious trioxide (1.3 mg/1 arsenic) in
a sodium arsenite solution. Concentrations of 4 mg/1 sodium arsenite
(2.3 mg/1 as arsenic) in confined outdoor pools have been found to reduce
survival and growth of fish and to reduce bottom fauna and plankton
populations (Gilderhus, 1966).
Trivalent arsenic is highly toxic to invertebrates. Conversely,
pentavalent arsenic is of relatively low toxicity. In Lake iErie
water Daphnia were observed to exhibit initial symptoms of immobility
at 18 to 31 mg/1 of sodium arsenate or 4.3 to 7.5 mg/1 as arsenic
(Anderson, 1944, 1946). The lethal threshold of sodium arsenate for
minnows has been reported to be 234 mg/1 as arsenic at 16° C to 20° C
(Wilber, 1969).
Ambient arsenic concentrations in sea water are reported to be
accumulated by oysters and other molluscan shellfish (Sautet, et_ al_., 1964;
-------
Lowman, et ah, 1971). Wilber (1969) reported concentrations of
100 mg/kg in shellfish.
Beginning in 1926 and for many years thereafter, sodium arsenite
w
was used for the control of vascular aquatic plants in the public lakes
of Wisconsin (Mackenthun, 1950). Dosages up to 10 mg/1 of the white
arsenic equivalent were used depending on the physical character of the
area to be treated. Harm to fish life within the lakes or to fishing
was not noted.
Lueschow (1964) found that Wisconsin lakes naturally may contain
10 ug/1 of arsenic. Pewaukee Lake, Wisconsin, a lake of 2300 acres,
received 215,174 pounds of arsenic for aquatic vegetation control in a
14-year period beginning in 1950. From 1959 to 1964, arsenic concen-
trations in the lake's outlet reached a maximum of 463 ug/1 arsenic and
consistently were above 100 ug/1. The mean arsenic concentration in
lake bottom muds was 208 ug/g on a dry weight basis. A single sample
of Cladophora sp. collected from Pewaukee Lake in 1962 contained 1258
ug/g arsenic (dry weight). Fresh shoots of mature Nlyri ophyll urn sp.
stems contained 228 and 561 ug/g arsenic (dry weight).
Benthic invertebrates, Chaoborus punctipenm's and Tendipedidae,
were present in Pewaukee Lake in populations that reached 288 per square
foot. Lueschow concluded that the trivalent inorganic arsenic was 10
to 15 times more toxic to Tendipedes plumosus than the pentavalent form.
His studies indicate that the trivalent inorganic arsenic is converted
to pentavalent arsenic within 30 days and that long-term survival of
-------
typical benthic organisms would be normal at concentrations as high
as 1,920 ug/g arsenic (dry weight) in lake muds.
The data cited above indicate that freshwater fish-food
organisms are adversely affected by concentrations of arsenic as low
as 1.3 mg/1, and the mobility of the freshwater crustacean Daphnia
is impaired by a concentration of arsenic as low as 4.3 mg/1. However,
these data are not considered to be sufficient to recommend any numerical
criterion for freshwater aquatic life. Such data as do exist indicate
that the 50 ug/1 criterion established for domestic water supplies should
be protective of aquatic life.
Rasmussen and Henry (1965) found that arsenic at 0.5 mg/1 in
nutrient solutions produced toxicity symptoms in seedlings of the
pineapple and orange. Below this concentration, no symptoms of
toxicity were found. Clements and Heggeness (1939) reported that
0.5 mg/1 arsenic as arsenite in nutrient solutions produced an 80
percent yield reduction in tomatoes. Following a review of available
literature the National Academy of Sciences (NAS, 1974) suggests that
a concentration of 100 ug/1 could be used for 100 years on sandy
soils, and a concentration of 2 mg/1 used for a period of 20 years
or 0.5 mg/1 used for 100 years on clayey soils with an adequate
margin of safety. Because of these factors, a criterion of 100 ug/1 for
the irrigation of crops is recommended. The herbicidal properties of arsenic
in water to aquatic vegetation.-also are recognized. Although data are not
sufficiently precise to recommend a criterion, such data as do exist
indicate that 100 ug/1 would be protective of aquatic vegetation.
30
-------
REFERENCES CITED
Anderson, B.G., 1944. The toxicity thresholds of various substances found in
industrial wastes as determined by the use of Daphnia magna. Sewage Works
J. 1616: 1156.
Anderson, B.G., 1946. The toxicity thresholds of various sodium salts determined
by the use of Daphnia magna. Sewage Works J. 18:82.
Borgono, J.M. and R. Grieber, 1972. Epidemiological study of arsenicism in
the city of Antofagasta. Proceedings of the University of Missouri's
Fifth Annual Conference on Trace Substances in Environmental Health (in press),
Boutwell, R.K., 1963. Feed additives: a careinogenicity evaluation of
potassium arsenite and arsanilic acid. Jour. Agr. Food Chem., 11: 381.
Chen, K.P. and H. Wu, 1962. Epidemiological studies on blackfoot disease:
II: A study of source of drinking water in relation to the disease. Jour.
Formosa Med. Assn., 61: 611.
Clements, H.F. and H.G. Heggeness, 1939. Arsenic toxicity to plants. Hawaii
Agr. Exp. Sta. Report, 1940: 77.
Craun, G. and L.J. McCabe, 1971. Waterborne disease outbreaks, 1961-1970.,
Presented at the annual meeting of the Amer. Water Works Assn. (June).
3-1
-------
DIPalma, J.R., 1965. Drill's pharmacology in medicine, 3rd ed. McGraw-Hill
Book Co., New York, p. 860.
DuBois, K.P. and E.M.K. Ceiling, 1959. Textbook of toxicology. Oxford
University Press, New York, p. 132.
Ducoff, H.S., et,aj[., 1948. Biological studies with arsenic (A8, 76). II.
Excretion and tissue localization. Proc. Soc. Exp. B1ol. Med. 69: 548.
Ferguson, J.F. and J. Gavis, 1972. A review of the arsenic cycle in
natural waters. Water Research, 6: 1259.
Frost, D.V., 1967. Arsenicals in biology—retrospect and prospect.
Fed. Amer. Soc. for Experimental Biol., 26:194.
Gilderhus, P.A., 1966. Some effects of sublethal concentrations of sodium
arsenite on bluegills and the aquatic environment. Trans. Amer. Fish.
Soc., 95: 289.
Goodman, L.S. and A.Z. Gilman, eds., 1965. The pharmacological basis of
therapeutics, 3rd ed. The McMillan Co., New York. p. 944.
Goldsmith, J.R., et_ al_., 1972. Evaluation of health implications of elevated
arsenic in well water. Water Research, 6: 1133.
Heuper. W.C. and W.W. Payne, 1963. Carcinogenic effects of adsorbates of
raw and finished water supplies. Amer. Jour. Clin. Path., 39: 475.
Kopp, J.F., 1969. The occurrence of trace elements in water, In: Proceedings
of the Third Annual Conference on Trace Substances in Environmental Health,
edited by D.D. Hemphill, University of Missouri, Columbia, p. 59.
-------
Lowman, F.G., e_t aK, 1971. Accumulation and redistribution of radlonuclldes
by marine organisms, In: Radioactivity in the Marine Environment,
National Academy of Sciences, Washington, D.C., p. 161.
Lueschow, L.A., 1964. The effects of arsenic trloxlde used 1n aquatic weed
control operations on selected aspects of the b1o-env1ronment. M.S. Thesis,
Univ. of Wisconsin, Madison, Wisconson.
flackenthun, K.M., 1950. Aquatic weed control with sodium arsenlte. Sew.
and Ind. Wastes, 22: 1062.
McCabe, L.J., et. al_., 1970. Survey of community water supply systems.
Jour. Amer. Water Works Assn., 62: 670.
Merck Index, 1968. An encyclopedia of chemicals and drugs, 8th ed.
Merck and Co., Inc., Rahway, N. J,
Musil, J. and V. Dejmal, 1957. Experimental and clinical administration
of radioarsenic (As76), Casopis.Lekari Ceskych, 96: 1543.
National Academy of Sciences, National Academy of Engineering, 1974. Water
quality criteria, 1972. U.S. Govt. Print. Office, Washington, D. C.
Paris, J.A., 1820. Pharmacologia: Comprehending the art of prescribing upon
fixed and scientific principles together with the history of medicinal
substances, 3rd ed. W. Philips, London, p. 132.
-------
Rasmussen, G.K. and W.H. Henry, 1965. Effects of arsenic on the growth of
pineapple and orange seedlings in sand and solution nutrient cultures.
Citrus Ind., 46: 22.
Sautet, J., ert a_l_.. 1964. Contribution to the study of the biological
fixation and elimination of arsenic by Mytilus edulis, Second Note.
Ann. Med. Leg. (Paris), 44: 466.
Schroeder, H.A. and J.J. Balassa, 1966. Abnormal trace metals in man:
Arsenic. Jour. Chronic 01s., 19: 85.
Smith, H., 1967. The distribution of antimony, arsenic, copper
and zinc in human tissue. J. Forensic Sci. Soc. , 7:97.
Soil man, T.H., 1957. A manual of pharmacology and its application to
therapeutics and toxicology, 8th ed. W. B. Sanders Co., Philadelphia, a.
Sommers, S.C. and R.G. McManus, 1953. Multiple arsenical cancers of skin
and internal organs. Cancer, 6: 347.
Surber, E.W. and L.O. Meehan, 1931. Lethal concentrations of arsenic for
certain aquatic organisms. Trans. Amer. Fish. Soc., 61: 225.
Trelles, R.A., e_t al_., 197.0. El problema sanitario de la
aguas destinadas a la bebida humana, con contenidos elevados de arsenico,
vanadio, y flour. Saneamiento, 34(217: 31-80).
Tseng, W.P., et^ al_., 1968. Prevalence of skin cancer in an endemic area of
chronic arsenicism in Taiwan. Jour. Nat. Cancer Inst., 40: 454.
-------
Wilber, C.G., 1969. The biological aspects of water pollution.
Charles C. Thomas, Publisher, Springfield, Illinois.
Yeh, S., et al., 1968. Arsenical cancer of the skin. Cancer, 21:312.
-------
BARIUM
CRITERION:
1 mg/1 for domestic water supply (health).
INTRODUCTION;
Barium is a yellowish-white metal of the alkaline earth group. It occurs
in nature chiefly as ba'rite, BaS04, and witherite, BaCOs, both of which are
highly insoluble salts. The metal is stable in dry air, but readily oxidized
by humid air or water.
Many of the salts of barium are soluble in both water and acid, and
soluble barium salts are reported to be poisonous (Lange, 196J; NAS, 1974).
However, barium ions generally are thought to be rapidly precipitated or removed
from solution by adsorption 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 appli-
cations including the metallurgic, paint, glass and electronics industries,
as well as for medicinal purposes.
RATIONALE:
Concentrations of barium in domestic drinking water supplies generally range
from less than 0.6 ug/1 to approximately 10 ug/1 with upper limits in a few midwestern
and western states ranging from 100 to 3,000 ug/1 (PHS, 1962/1963; Katz, 1970;
Little, 1971). Barium enters the body primarily through air and water, since
-------
appreciable amounts are not contained in foods (MAS, 1974).
The fatal dose of barium for man is reported to be 550 to 600 mg.
Ingestion of soluble barium compounds may also result in effects on the
gastrointestinal tract causing vomiting and diarrhea and on the central
nervous system causing violent tonic and clonic spasms followed in some
cases by paralysis (Browning, 1961; and 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 (Sollmann,
1957). By constricting blood vessels, barium may cause an increase in blood
pressure. On the other hand, it is not likely that barium accumulates in
the bone, muscle, kidney, or other tissues because it is readily excreted
(Browning, 1961; McKee and Wolf, 1963).
Stokinger and Woodward (1958) developed a safe concentration for
barium in drinking water based on the limiting values for industrial
atmospheres, an estimate of the amount absorbed into the blood stream,
and daily consumption of two liters of water. From these factors they
arrived at a limiting concentration of 2 mg/1 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/1 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/1 before
toxicity to aquatic life would be expected. In most natural waters, there
is sufficient sulfate or carbonate to precipitate the barium present in
the water as a virtually insoluble, non-toxic compound. Recognizing that
the physical and chemical properties of barium generally will preclude the
existence of the toxic soluble form under usual marine and freshwater
conditions, a restrictive criteria for aquatic life appears unwarranted.
37
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REFERENCES CITED:
Browning, E., 1961. Toxicity of Industrial metals. Butterworths,
London, England.
Katz, M., e£al_., 1970. Effects of pollution on fish life, heavy rnetals,
Annual literature review, Jour. Water Poll. Cont. Fed., 42:987.
Lange, N.A., 1961. Handbook of chemistry. McGraw-Hill, Tenth Ed.
Little, A.D., 1971. Inorganic chemical pollution of freshwater. Water
Quality Data Book, Vol. 2. Environmental Protection Agency, 18010 DPV,
pp. 24-26.
McKee, J.E. and W.W. Wolf, 1963. Water quality criteria, California State
Water Resources Control Board, Pub. No. 3-A.
National Academy of Sciences, National Academy of Engineering, 1974. Water
quality criteria, 1972. U.S. Government Printing Office, Washington, D.C.
Patty, F.A. Industrial Hygiene and Toxicology, Vol. II (New York: Wiley, pp 998-
1002, 1962) cited in U.S. Department of Health, Education and Welfare, 1969.
Sollmann, T.H., 1957. A manual of pharmacology and its applications to
therapeutics and toxicology. 8th ed. W.B. Saunders Co., Phil a. Pa.
Stokinger, H.E. and R.L. Woodward, 1958. Toxicologic methods for
establishing drinking water standards. Jour. Amer. Water Works Assn., 50:515.
U.S. Department of Health, Education and Welfare, 1969. Preliminary Air
Pollution Survey of Barium and Its Compounds, A Literature Review. National
Air Pollution Control Administration Publication No. APT D 69-28, Raleigh, NC.
U.S. Public Health Service, 1962/63. Drinking water quality of selected
interstate carrier water supplies. U.S. Dept. of Health, Education and
Welfare, Wash., D.C.
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BERYLLIUM
CRITERIA:
11 ug/1 for the protection of aquatic life
in soft fresh water; *
1,100 ug/1 for the protection of aquatic life
in hard fresh water; *
100 ug/1 for continuous irrigation on all
soils; except
500 ug/1 for irrigation on neutral
to alkaline fine-textured soils.
INTRODUCTION:
Beryllium is not likely to occur at significantly toxic levels
in ambient natural waters (McKee and Wolf, 1963). Although the chloride
and nitrate salts of beryllium are very water soluble, and the sulfate
is moderately so, the carbonate and hydroxide are almost insoluble in
cold water (Lange, 1961). Kopp and Kroner (1967) reported that for 1,577
surface water samples collected at 130 sampling points in the United
States, 85 samples (5.4 percent) contained from 0.01 to 1.22 ug/1 with
a mean of 0.19 ug/1 beryllium. The concentration of beryllium in sea
water is 6 x 10"4 ug/1 (Goldberg, ejt aJL, 1971).
RATIONALE:
The major human toxic hazard potential of beryllium is via the
inhalation of beryllium-containing fumes and dusts that might emanate
from processing and fabrication operations. Beryllium could enter
waters in effluents from certain metallurgical plants (NAS, 1974).
*See criteria for Hardness (p. 147). The beryllium concentration that
will be protective of a given aquatic ecosystem can be obtained by
conducting flow-through bioassays using ambient water and native species
of fish and invertebrates.
-------
Contact dermatitis, characterized by itching and reddened, elevated,
or fluid-accumulated lesions, which appear particularly on the exposed
surfaces of the body, may occur either on an allergic basis or from
primary irritation following contact with soluble beryllium salts (Van
Ordstrand, ejt a]_., 1945; McCord, 1951). A latent period is occasionally
noted, indicating the development of delayed hypersensitivity (NIOSH, 1972),
Ocular effects may occur as inflammation of the conjunctiva in
"splash burn" or in association with contact dermatitis (Van Ordstrand,
e_t aj_., 1945). Splashes may also cause corneal burns closely resembling
those produced by acids and alkalies, and fluid accumulation and reddening
around the eye*socket are frequently noted (NIOSH, 1972).
Beryllium is demonstrably toxic by most routes of administration
(NIOSH, 1972), but its oral toxicity is notably different from that by
other routes. The sulfate, for example, while highly toxic by all
other routes at a single dose level, is practically non-toxic by mouth
at a level several thousand-fold greater by multiple dose (Reeves,
1965; Stokinger and Stroud, 1951).
Tarzwell and Henderson (1960) obtained 96-hour LC5Q values ranging
from 0.15 mg/1 beryllium (vjfeen tested as the nitrate and chloride) to
0.2 mg/1 beryllium (when tested as the sulfate) for fathead minnows in soft
water (20 mg/1 CaCX>3/ total alkalinity of 18 mg/1, and pH of 7.4),
and from 11 mg/1 beryllium (as sulfate) to 20 mg/1 beryllium (as nitrate)
for fathead minnows in hard water (400 mg/1 CaCO , total alkalinity of
-------
360 mg/1, and pH of 8.2). For bluegill they obtained 96-hour LC5Q values
of 1.3 mg/1 beryllium in soft water and 12 mg/1 beryllium in hard water
(both as the sulfate.)
Slonim (1973) obtained 96-hour LC_Q static bioassay values of
0.19 mg/1 beryllium for the comon guppy, Efe6|pilia reticulata,
in soft water with a hardness of 20 to 25 mg/1 as CaCCL, total
alkalinity of 16 to 18 mg/1, and pH of 6.3 to 6.5, and 20.3 mg/1
beryllium in hard water with a hardness of 400 to 500 mg/1,
alkalinity of 185 to 230 mg/1, and pH of 7.8 to 8.2.
Slonim and Slonim (1973) studied the influence of water hardness
on the toxicity of beryllium sulfate to the common guppy, Poecilia
reticulata, by simultaneously conducting static bioassays at 400, 275,
15.0, and 22 mg/1 hardness as CaCC^. The 96-hour 1X50 values were
20.0, 13.7, 6.1, and 0.16 mg/1, respectively. In a water hardness of
22 mg/1, 80 percent of the test fish survived 0.1 mg/1 beryllium
for 96 hours-, whereas 100 percent survived control conditions. In a
water hardness of 150 mg/1, 70 percent survived 5 mg/1 beryllium for
96 hours and 100 percent survived 2.5 mg/1 beryllium.
Slonim and Ray (1975) studied the influence of water hardness on
the toxicity of beryllium sulfate to two species of salamander larvae
by conducting static bioassays at hardness levels of 400 mg/1 and 20
to 25 mg/1 as CaCOg. , In hard water all of the test organisms survived
exposure to 10 mg/1 beryllium throughout the exposure period, whereas
there was a significant decline in survival with time in soft water.
-------
The 96-hour LCgo values were 26.3 mg/1 beryllium in hard water, and
4.7 mg/1 beryllium in soft water.
Based on the fathead minnow and bluegill data of Tarzwell and
Henderson (1960), the observations of Slonim and Ray (1975) on salamander
larva survival, and the observations of Slonim and Slonim (1973) on
guppy survival, the criterion for the protection of aquatic biota is
established at 1.1 mg/1 beryllium in hard fresh water. This value
assumes an application factor of 0.1 of the 96-hour LC5n value for
fathead minnows. For soft fresh water, in view of the reported approxi-
mate 100-fold increase in acute fish toxicity over that found in hard
water, the criterion for the protection of aquatic biota is set at 0.011
mg/1 beryllium.
Beryllium has been reported to be concentrated 1000 times in
marine organisms (Goldberg, ejt aK, 1971). The average concentration
factors for marine benthic algae, phytoplankton, and zooplankton also
have been reported as 110, 1000, and 15 mg/1, respectively (Lowman,
e_t al_., 1971). Riley and Roth (1971) reported levels of 1.1 to 18 mg/1
beryllium for various species of marine algae. The 96-hour TLm value
for beryllium resulting from tests performed on the killifish, Fundulus
heteroclitus. was 41 mg/1 (Jackim, e_t afL, 1970). These data do not
represent an adequate base upon which to establish a marine criterion.
Beryllium has been shown to inhibit photosynthesis in terrestrial
plants (Bollard and Butler, 1966), and to be toxic to some varieties
-------
of citrus seedlings at 2.5 mg/1 beryllium (Haas, 1932). Romney, et al_. (1962)
found that beryllium at 0.5 mg/1 in nutrient solutions reduced the
growth of bush beans, and Romney and Chi 1dress (1965) found that con-
centrations of 2 mg/1 or greater in nutrient solutions reduced the growth
of tomatoes, peas, soybeans, lettuce, and alfalfa plants. Additions of
soluble beryllium salts at levels equivalent to 4 percent of the cation-
adsorption capacity of two acid soils reduced the yields of ladino
clover; beryllium carbonate and beryllium oxide at the same levels did
not reduce yields, suggesting that beryllium in calcareous soils might
•
be less active than in acid soils. Williams and LeRiche (1968) found
that beryllium at 2 mg/1 in nutrient solutions was toxic to mustard,
whereas 5 mg/1 was required for growth reduction in kale. In view of
its toxicity in nutrient solutions in acid soils, the criteria for
berylli*111 in irrigation waters are 0.10 mg/1 fear use on all soil?-
exoept 0.50 mg/1 for use on neutral to alkaline fine-textured soils.
-------
REFERENCES CITED:
Bollard, E.G. and G.W. Butler, 1966. Mineral nutrition of plants.
Ann. Rev. Plant Physiol., 17:77.
Goldberg, E.D., et^ al_., 1971. Marine chemistry. In: Radioactivity
in the marine environment. National Academy of Sciences, Washington,
D. C., p. 137.
Haas, A.R.C., 1932. Nutritional aspects in mottleleaf and other
physiological diseases of citrus. Hilgardia, 6:483.
Jackim, E., ejt al_., 1970. Effects of metal poisoning for five liver
enzymes in the killifish (Fundulus heteroclitus). Jour. Fish. Res.
Bd. Can., 27:383.
Kopp, J.F. and R.C. Kroner, 1967. Trace metals in waters of the United
States. U.S. Department of the Interior, FWPCA, Cincinnati, Ohio.
Lange, N.A., Ed., 1961. Handbook of chemistry, 10th ed. McGraw-Hill
Book Co., New York.
Lowman, F.G., et al_., 1971. Accumulation and redistribution of
radionuclides by marine organisms. In: Radioactivity in the marine
environment. National Academy of Sciences, Washington, D. C., p. 161.
McCord, C.P., 1951. Beryllium as a sensitizing agent. Ind. Med. Surg.,
20:336.
McKee, J.E. and H.W. Wolf, 1963. Water quality criteria. State
Water Quality Control Board, Sacramento, Calif., Pub. 3-A.
HH
-------
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972. U.S. Government Printing Office,
Washington, D. C.
National Institute for Occupational Safety and Health, 1972. Occupational
exposure to beryllium. U.S. Department of Health, Education and Welfare,
Health Services and Mental Health Admin., National Institute for
Occupational Safety and Health. U.S. Government Printing Office,
Washington, D. C.
Reeves, A.L., 1965. Absorption of beryllium from the gastrointestinal
tract. A.M.A. Arch. Envir. Health, 11:209.
Riley, J.P. and I. Roth, 1971. The distribution of trace elements in
species of phytoplankton grown in culture. Jour. Mar. Biol. Assn.
U.K., 51:63.
Romney, E.M., et al_., 1962. Beryllium and the growth of bush beans.
Science, 135:786.
Romney, E.M. and J.D. Childress, 1965. Effects of beryllium in plants
and soil. Soil Sci., 100:210.
Slonim, A.R., 1973. Adute toxicity of beryllium sulfate to the common
guppy. Jour. Water Poll. Cont. Fed., 45:2110.
Slonim, A.R. and E.E. Ray, 1975. Acute toxicity of beryllium sulfate
to salamander larvae (Ambystoma spp.). Bull. Envir. Contam. Toxicol.,
13:307.
-------
Slonim, C.B. and A.R. Slonim, 1973. Effect of water hardness on the
tolerance of the guppy to beryllium sulfate. Bull. Envir. Contam.
Toxicol., 10:295.
Stokinger, H.E. and C.A. Stroud, 1951. Anemia in acute experimental
beryllium poisoning. Jour. Lab. Clin. Med., 38:173.
Tarzwell, C.M. and C. Henderson, 1960. Toxicity of less common metals
to fishes. Ind. Wastes, 5:12.
Van Ordstrand, H.S., e_t aK, 1945. Beryllium poisoning. Jour.,
Amer. Med. Assn., 129:1084.
Williams, R.J.B. and H.H. LeRiche, 1968. The effects of traces of
beryllium on the growth of kale, grass, and mustard. Plant Soil, 29:317.
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BORON
CRITERION;
750 ug/1 for long-term Irrigation on sensitive crops.
INTRODUCTION:
Boron is not found In Its elemental form in nature; it 1s 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. Berates are used as "burnable" poisons.
RATIONALE;
Boron is an essential element for growth of plants but there 1s
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/1; the mean value was 0.1 mg/1 (Kopp and Kroner,
1967). Ground waters could contain substantially higher concentrations
at certain places. The concentration in sea water is reported as 4.5 mg/1
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 add at 20° C
for 6 hours was reported to be 18,000 to 19,000 mg/1 in distilled
water and 19,000 to 19,500 mg/1 in hard water (Le Clerc and Devlaminck,
1955; Le Clerc, 1960).
-------
In the dairy cow, 16 to 20' g/day of boric add for 40 days
produced no ill effects (McKee and Wolf, 1963).
Sensitive crops have shown toxic effects at 1000 ug/1 or less
of boron (Richards, 1954). Bradford (1966), in a review of boron
deficiencies and toxicities, stated that when the boron concentration
1n irrigation waters was greater than 0.75 mg/1, 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 mg/1 boron might be used for some time without injury
to sensitive plants. Th*'criterion of 750 ug/1 1s thought to protect
sensitive crops during long-term irrigation.
-------
REFERENCES CITED:
Biggar, J.W. and M. Fireman, 1960. Boron absorption and release by
soils. Soil Sci. Soc. Amer. Proc., 24:115.
Bradford, G.R., 1966. Boron [toxicity, indicator plants], in diagnostic
criteria for plants and soils. H.D. Chapman, Ed., University of
California, Division of Agricultural Science, Berkeley, p. 33.
Le Clerc, E., 1960. The self-purification of streams and the relation-
ship between chemical and biological tests. Proc. Second Symposium on
Treatment of Waste Waters, Pergamon Press, London, England, p. 281.
Le Clerc, E. and F. Devlaminck, 1955. Fish toxicity tests and water
quality. Bull, de Beige Condument Eaux., 28:11.
Kopp J.F. and R.C. Kroner, 1967. Trace metals in waters of the United
States. U.S. Dept. of Interior, Federal Water Pollution Control Admin.,
Cincinnati, Ohio.
McKee, J.E. and H.W. Wolf, 1963. Water quality criteria. State Water
Quality Control Board, Sacramento, Calif., Pub. 3-A.
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972. U.S. Government Printing Office,
Washington, D. C.
Richards, L.A. (Ed.), 1954. Diagnosis and improvement of saline and
alkali soils. Agriculture Handbook No. 60, U.S. Government Printing
Office, Washington,".D. C.
-------
CADMIUM
10 ug/1 for domestic water supply (health) .
Aquatic Life:
Fresh Water
Soft Water* Hard Water*
0.4 ug/1 1.2 ug/1 for cladocerans anrl salmnniri fishes;
4.0 ug/1 12.0 ug/1 for other, less sensitive, aouatic life.
INTRODUCTION;
Cadmium is a soft, white, easily fusible metal similar to zinc and
lead in many properties and readily soluble in mineral acids. Biologically,
cadmium is a nonessential, nonbeneficial element recognized to be of
high toxic potential. It is deposited and accumulated in various body
tissues and is found in varying concentrations throughout all areas
where man lives. Within the past two decades industrial production
and use of the metal has increased. Concomitantly, there have been
incidences of acute cases of clinically identifiable cadmiosis. Cadmium
may function in or may be an etiological factor for various human
pathological processes including testicular tumors, renal dysfunction,
hypertension, arteriosclerosis, growth inhibition, chronic diseases of
old age, and cancer.
*See Criteria for Hardness (p. 147). The cadmium concentration that will
be protective of a given aquatic ecosystem can be obtained by conducting
flow-through bioassays using ambient water and native species of fish and
invertebrates.
-------
Cadmium occurs in nature chiefly as a sulfide salt, frequently in
association with zinc and lead ores. Accumulations of cadmium in soils
in the vicinity of mines and smelters may result in high local concen-
trations in nearby waters. The salts of the metal also may occur in
wastes from electroplating plants, pigment works, textile and chemical
industries. Seepage of cadmium from electroplating plants has resulted
in groundwater cadmium concentrations of 0.01 to 3.2 mg/1 (Lieber and
Welsch, 1954). Kopp and Kroner (1967) on one occasion reported 120 ug/1
dissolved cadmium in the Cuyahoga River at Cleveland, Ohio. However,
dissolved cadmium was found in less than 3 percent of 1,577 water samples
examined for the United States, with a mean of slightly under 10 ug/1.
Most fresh waters contain less than 1 ug/1 cadmium and most analyses of
sea water indicate an average concentration of about 0.15 ug/1 (Fleischer,
et al., 1974).
RATIONALE!
Cadmium has been shown to be toxic to man when ingested or inhaled.
Exposure by the former route causes symptoms resembling food poisoning.
A group of school children became ill after eating popsicles containing
13 to 15 mg/1 cadmium (Frant and Kleeman, 1941), and this level, equivalent
to 1.3 to 3.0 mg of cadmium ingested, commonly is considered the emetic
threshold concentration.
In a specially designed study, five groups of rats were exposed to
drinking water containing cadmium concentrations of 0.1 mg/1 to
10 mg/1. Although no visible toxic effects were noted, the content
of cadmium in the kidney and liver increased in direct proportion
-------
to the dose at all levels of exposure. At the end of one year, tissue
concentrations were approximately double those found after six months
(Ginn and Volker, 1944). Later work confirmed that virtually no absorbed
cadmium was eliminated (Decker, et a^., 1958).
Drinking water containing excessive cadmium led to the occurrence
of itai-itai disease among the Japanese (Kobayashi, 1970). Within 15
years 200 cases of this disease were recorded in the Jintsu River Valley,
half of which resulted in death. The disease is characterized by
rheumatic symptoms with intense pain in the bones caused by a loss of
bone minerals with the bones becoming as flexible as soft tissues. Yamagata
and Shigematsu (1970) estimated that a daily intake of 600 ug cadmium would
not produce itai-itai disease in an endemic area.
Studies in animals (Ferm, et al_., 1969; Chernoff, 1973) as well as
observations in human fetuses (Chaube, et al^., 1973; and Friberg, et al./
1974) indicate that the placenta is not a complete barrier against the
transfer of cadmium. About 6 percent of the cadmium that enters the
stomach via food or drink is absorbed into the body tissues and about
30 percent of the cadmium inhaled into the lungs by breathing and smoking
is absorbed (EPA, In Press). Presently there are no known physiological
needs for cadmium and no mechanism by which the body maintains cadmium at
a constant safe level. Once absorbed, it is stored largely in the kidneys
and liver and is excreted at an extremely slow rate.
C2.
-------
Chronic kidney disease (renal tubular dysfunction) will begin to
occur in an individual when the cadmium accumulated in 'the kidneys reaches
a critical concentration. While the critical concentration varies from
one individual to another, the threshold of observed effect is about 200 ppm
of cadmium in the renal cortex (Friberg, et al., 1974). Individuals have
been found with several times this level without evidence of kidney disease.
At a 10 ug/1 criterion, as recommended in drinking water, the maximum
daily intake of cadmium would not exceed 20 ug from water, assuming a
2-liter daily consumption. The daily intake from other sources is up to
60 ug in the population of the United States (Schroeder, et al., 1961).
Pickering and Cast (1972) conducted two separate flow-through tests
on the chronic toxicity of cadmium to the fathead minnow, Pimephales
promelas, using water of 202 mg/1 hardness, 157 mg/1 alkalinity, and
7.7 pH. Five cadmium concentrations from 4 to 350 ug/1 were delivered
to the exposure chambers in each test over the life history of the fish.
A concentration of 57 ug/1 cadmium decreased survival of developing embryos.
At levels from 4.5 to 37 ug/1 no adverse effect on survival, growth, or
reproduction was found. Eaton (1974b) exposed bluegill sunfish, Lepomis
macrochirus, to five cadmium concentrations ranging from 31 to 2,140 ug/1
for 11 months in a flow-through system using water of the same hardness as
above. Nine of the 18 adult bluegill sunfish exposed to 80 ug/1 died
by the end of the test, while all of those exposed to 31 ug/1 and the
control survived. Although at 80 ug/1 cadmium the hatchability of eggs
was not measurably affected, the survival and growth of the resulting
larvae were severely reduced after 60 days. Larvae exposed to 31 ug/1
cadmium survived and grew about as well as the control fish. Sixty days
-------
after hatching in. hard water, growth and survival of channel catfish fry,
Ictalurus punctatus, was reduced significantly at a cadmium concentration
of 17 ug/1 but not at 12 ug/1 (Eaton, 1974a). Thus, in hard water, a
criterion of 12 ug/1 cadmium represents a demonstrated no-effect level
for catfish and therefore was chosen to protect non-salmonid freshwater
fish species. No data are available upon which to base an acceptable
concentration for the chronic exposure of sensitive vertebrates or
invertebrates in hard water, therefore a criterion is proposed which
is reduced from the less sensitive aquatic life value by the same
factor (0.1) as the criterion for soft water, i.e. 1.2 ug/1 for salmonids
and invertebrates.
Spehar (1974) reported on chronic toxicity tests with cadmium using
a topminnow native to Florida in water with a hardness of 41 to 45 mg/1
as CaCO , alkalinity of 38 to 43 mg/1, and a pH of 7.4. There was a
-------
When embryos and larvae-juveniles of brook trout, Salvelinus fontinalis;
brown trout, Saline trutta; lake trout, Salvelinus namaycush; northern pike,
Esox lucius; white suckers, Catastonms oommersoni; smallirouth baiss,
Micropterus dolomieui; and coho salmon, Oncorhynchus kisutch, were exposed
to cadmium concentrations ranging from 0.4 to 100 ug/1 in soft water, all
suffered a reduction in standing crop at 12 ug/1. Standing crop) survival
times and weights were virtually equal among the four species when exposed
to 4 ug/1; 1.2 ug/1 was the highest cadmium concentration not causing
a reduction among the other three salmonid species after 30 to 120 days
of exposure (Eaton, 1974a).
Tests in water with 23 mg/1 hardness, 18 mg/1 alkalinity, and a pH
of 7.3 in a flow-through system indicated a 96-hour LC50 of 2.0 ug/1 for
the initial feeding stage of chincok salmon, Qnoorhynchus tshawytscha,
and 0.92 ug/1 for 5-tnonth-old steelhead trout, Salmo gairdneri. A
criterion of 0.4 ug/1 for cadmium is believed to offer protection to the
cladocerans and salmonids in soft water (Eaton, 1974a).
Anderson, et al. , (1975) , determined the 10-day LC50 for
the midge Tanytarsus dissimilis to be 3.4 ug/1 in a flow-through system, soft
water (48 mg/1) bioassay involving at least one molt. A concentration
of 3.1 ug/1 retarded growth but 1.9 ug/1 elicited no obvious effect.
Biesinger and Christensen (1972) measured the toxicity of cadmium to
Daphnia magna during an entire life cycle in test water with a hardness
of 45 mg/1, alkalinity of 42 mg/1, and a pH of 7.7. It was found that
-------
50 percent of the daphnids exposed to cadmium concentrations of 5 ug/1
were killed in three weeks. The production of young was reduced by
50 percent compared to the controls in a cadmium concentration of 0.7 ug/1.
Several invertebrate species have been found much less sensitive to
cadmium in acute tests than in the midge and cladoceran exposures cited above.
(Rehwoldt, et aJL., 1973; Thorp and Lake, 1974; Warnick and Bell,
1969). These data support a criterion for some invertebrates in soft
water that is identical to that for salmonids.
Steward and Pesche, as reported by Eisler (1974), stated that for
grass shrimp, Palaemonetes pugio, which were subjected to cadmium chloride
in flowing sea water, 69 percent died in 43 days in 500 ug/1 cadmium and
10 percent died in a 250 ug/1 cadmium solution. Hermit crabs, Pagurus
longicarpus, all died in a 250 ug/1 cadmium solution, and 30 percent
died in 43 days in a 120 ug/1 solution. In 63 days, 40 percent died
in a 60 ug/1 cadmium solution. All of the controls survived.
Eisler (1971) found that the 96-hour LC50 for three species of
marine decapod crustaceans ranged between 320 and 420 ug/1 at 20° C and
20 mg/1 alkalinity. The 96-hour I£25 was 180 ug/1 for the hermit
crab, Pagurus longicarpus, and the grass shrimp, Palaemonetes vulgaris,
and 80 ug/1 for the sand shrimp, Crangon septernspinosa.
Zaroogian, as reported by Eisler (1974jT), states that adult oysters,
Crassostred virginica, exposed to 10 ug/1 cadmium between April
1973 and August 1973 accumulated 18,000 ugAg of cadmium in wet whole
-------
neat, which exceeds the human enetic threshold of 13,000 to 15,000 ug/kg.
Oysters retained virtually all of the accumulated cadmium for at least
several months and some histopathology was evident. Under natural
conditions, significantly greater numbers of larvae from cadmium-stressed
oysters failed to develop when compared to controls after 48 or 72 hours.
A criterion of one-half of the level at which oysters accumulate cadmium
in excess of the human emetic threshold (i.e. 5 ug/1) is believed to
provide protection for consumers of oysters.
Data by Page, et al. (In Press), show that the yield of beans, beets,
and turnips was reduced about 25 percent by 0.10 mg/1 cadmium in nutrient
solutions; whereas cabbage and barley yields decreased 20 to 50 percent
at 1.0 mg/1.
Yamagata and Shigematsu (1970) have demonstrated that foods cultured
on cadmium-polluted soils irrigated with cadmium-polluted water can
accumulate sufficient cadmium to be hazardous to humans who consume these
foods. Chaney (1973) suggests that it is not the cadmium concentration
per se in the soil that determines cadmium accumulation by plants and
that as long as the ratio of zinc to cadmium is 100 or greater, foods
will not accumulate hazardous concentrations of cadmium. Hie subject is
complex and additional research is needed to resolve potential, hazards
associated with the cadmium-zinc-soil-plant system.
-------
Fish and certain invertebrates have been found to be sensitive to
low levels of cadmium in water. Salmonids and cladocerans appear to
be the most sensitive among organisms tested. Increased hardness and/or
alkalinity have been demonstrated to decrease the toxicity of cadmium
in acute freshwater mortality tests, but may have less of an effect at
low cadmium levels. Edible marine organisms can concentrate cadmium
levels and become hazardous to be ultimate consumer. Lovman, et al. (1971)
reported a concentration factor of 1,000 for cadmium in fish muscle. The
criteria necessary to protect fish and other aquatic life are more stringent
than those necessary to protect a public water supply or other uses. Data
support a division of criteria for "hard" and "soft" water environments.
The data used to develop a criteria for cadmium were obtained over
a range of hardness, alkalinity and pH values. Their usefullness in
developing criteria will be limited to the range of physical parameters
for which experimental data are available. Interpolation of the present
data will be required to derive criter ie> for aquatic ecosystems which
do not approximate the experimental conditionsf
-------
REFERENCES CITED:
Anderson, R.A., eJrjsJy 1975. Survival and growth of
Tanytarsus dissimilis (Chroncrnidae) exposed to copper, cadmium,
zinc and nickel. National Environmental Research
Laboratory,Duluth, Minn., Quarterly Reports.
Benoit, D.A., ejt al_., Chronic
effects of cadmium on three generations of brook trout (Salvelinus
fontinalis), EPA Ecological Research Series, In Press.
Biesinger, K.E. and 6.M. Christensen, 1972. Effects of various metals
on survival, growth, reproduction, and metabolism of Daphnia magna.
Jour. Fish. Res. Bd. Can., 29:1691.
Chancy, R.L., 1973. Crop and food chain effects of toxic elements in
sludges and effluents. U.S. Environmental Protection Ayency, U.S.
Department of Agriculture, Universities Workshop, Champaign, 111. (July).
Chaube, s., et. aJL, 1973. Zinc and cadmium
in normal human embryos and fetuses. Arch. Environ. Health, 26:237.
Chernoff, N., 1973. Teratogenic effects of cadmium in rats. Teratology,
8(1):27-32.
Decker, I.E., et al_., 1958. Chronic toxicity studies. I: Cadmium
administered in drinking water to rats. A.M.A. Arch. Ind. Health, 18:228.
Eaton, J.G., 1974a. Testimony in the matter of proposed toxic pollutant
effluent standards for Aldrin-Dieldrin et al_. FWPCA (307), Docket No. 1.
Eaton, J.G., 1974b. Cadmium toxicity to the bluagill (Lepomis macrochirus
Rafinesque). Trans. Amer. Fish. Soc., 103:729.
5ft
-------
•Eisler, R., 1971. Cadmium poisoning in Fundulus heteroclitus (Pisces:
Cyprinodontidae) and other marine organisms. Jour. Fish. Res. Bd.
Can., 28:1225.
^Eisler, R., 1974. Testimony in the matter of proposed toxic pollutant
effluent standards for Aldrin-Dieldrin e_t al_. FWPCA (307) Docket No. 1,
i EPA, 1975. Scientific summary document on cadmium. (In Press)
(.Perm, V.H., et al., 1965. The permeability of the hamster
placenta to radioactive cadmium. J. Embryol. Exp. Morph.
22:107.
"Fleischer, M., et aJL, 1974. Environmental impact of cadmium: A
review by the-panel on hazardous trace substances. Env. Health
Perspectives, U.S. Government Printing Office, Washington, D.C., 7:253.
' Frant, S. and I. Kleeman, 1941. Cadmium "food poisoning." Jour.
Amer. Med. Assn., 17:86.
' Friberg, L., et al., 1974. Cadmium in the enviroment.
2nd E. CHC Press, Cleveland, pp 30-31.
^Ginn, J.T. and J.F. Volker, 1944. Effect of Cd and F on rats. Proc.
Soc. Exp. Biol. Med., 57:189.
^Kobayashi, J., 1970. Relation between the "itai-itai" disease and the
pollution of river water by cadmium from a mine. Fifth International
Water Pollution Research Conference, 1-25, 1-7.
-------
Kopp, J.F. and R.C. Kroner, 1967. Trace metals in waters of the United
States. U.S. Dept. of the Interior, Federal Water Pollution Control
Admin., Cincinnati, Ohio.
' Lieber, M. and W.E. Welsch, 1954. Contamination of ground water by
cadmium. Jour. Amer. Water Works Assn., 46:51.
i/Lowman, F.G., et al., 1971. Accumulation and redistribution of radio-
nuclides by marine organisms. In: Radioactivity in the Marine
Environment, National Academy of Sciences, Washington, D. C., p. 161.
" Page, A.L., e_t a/L (In Press). Cadmium absorption and growth of various
plant species as influenced by solution cadmium concentration. Jour.
of Environmental Quality.
'Pickering, Q.H. and M.H. Cast, 1972. Acute and chronic toxicity of
cadmium to the fathead minnow (Pimephales promelas). Jour. Fish.
Res. Bd. Can., 29:1099.
"Rehwoldt, R., et. al... 1973. The acute
toxicity of seme heavy metal ions toward benthic organisms. Bull.
Environ. Contam. and Tox., 10:291.
^Schroeder, H.A., et al_., 1961. Abnormal trace metals in man: Cadmium.
Jour. Chron. Dis., 14:236.
^Spehar, R., 1974. Cadmium and zinc toxicity to Jordanella floridae.
M.S. Thesis, University of Minnesota, Duluth.
-------
v Thorpe, V.J. and P.S. Lake/ 1974. TQxicity bioassays of cadmium on
selected freshwater invertebrates and the interaction of cadmium and
ainc on the freshwater shrimp, Paratya tasmaniensis, Riek. Aust.
Jour. Mar. Freshwater Research, 25:97.
i/Warnick, S.L. and H.L. Bell, 1969. The acute toxicity of some heavy
metals to different species of aquatic insects. Jour. Water Poll.
Control Fed., 41:280.
iVamagata, N. and I. Shigematsu, 1970. Cadmium pollution in perspective.
Bull. Inst. Public Health, 19:1.
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CHLORINE
CRITERIA:
Total residual chlorine:
2.0 ug/1 for salmonid fish ;
10.0 ug/1 for other freshwater and marine organisms.
INTRODUCTION:
Elemental chlorine is a greenish-yellow gas that is highly
soluble in water. It reacts readily with many inorganic substances
and all animal and plant tissues. The denaturing effect of chlorine
on animal and plant tissues forms the basis for its use as an effec-
tive water or wastewater disinfectant. When chlorine dissolves .in
water, it hydrolyzes according to the reaction: C12 + H20— -> HOC1 +
H+ +CL~. Unless the concentration of the chlorine solution Is above
1000 mg/1, all chlorine will be in the form. of HOC1 or its dissociated
ions H+ and OC1~. The HOC1 is a weak acid and is dissociated according
to the equation, HOC! t=, H+ +OCT.
The ratio between HOC1 and OC1~ is a function of the pH, with
96 percent HOC1 remaining at pH 6, 75 percent at pH 7, 22 percent at
pH 8, and 3 percent at pH 9. The relationship of HOC1 to pH 1s
significant as the undissociated form appears to be the bactericidal
agent in the use of chlorine for disinfection (Moore, 1951).
Chlorine is not a natural constituent of water. Free available
chlorine (HOC1 and OCT) and combined available chlorine (mono- and
fa!
-------
di-chloramines) appear transiently 1n surface or ground waters as a
result of disinfection of domestic sewage or from industrial processes
that use chlorine for bleaching operations or to control organisms that
grow in cooling water systems.
RATIONALE:
Chlorine, in the free available form reacts readily with
nitrogenous organic materials to form chloramines. These compounds
are toxic to fish. Chloramines have been shown to be slightly less
toxic to fish than free chlorine, but their toxicity is considered
to be close enough to free chlorine that differentiation is not
warranted (Merkens, 1958). Since the addition of chlorine or hypochlorites
to water containing nitrogenous materials rapidly forms chloramines,
toxicity in most waters is related to the chloramine concentration. The
toxicity to aquatic life of chlorine will depend upon the concentration
of total residual chlorine, which is the amount of free
chlorine plus chloramines. The persistence of chloramines is dependent
on the availability of material with a lower oxidation-reduction
potential.
In field studies in Maryland and Virginia, Tsai (1973) observed
that downstream from plants discharging chlorinated sewage effluents,
the total numbers of fish species were drastically reduced with the
stream bottom clear of aquatic organisms characteristically present
in unchlorinated wastewater discharges. No fish were found in water
with a chlorine residual above 0.37 mg/1 and the species diversity
-------
index reached zero at 0.25 mg/1. A 50 percent reduction in the
species diversity index occurred at 0.10 mg/1. Of the 45 species
of fish observed in the study areas, the brook trout and brown trout
were the most sensitive and were not found at residual chlorine levels
above about 0.02 mg/1. In studies of caged fish placed in waters
downstream from chlorinated wastewater discharges, the Michigan Depart-
ment of Natural Resources (1971) reported that 50 percent of the rainbow
trout died within 96 hours at residual chlorine concentrations of 0.014
to 0.029 mg/1. Some fish died as far as 0.8 miles (1.3 km) downstream
from the outfall.
Studies described by Brungs (1973) indicate that salmom'ds are
the most sensitive fish to chlorine. A residual chlorine concentration
of 0.006 mg/1 was lethal to trout fry in two days (Coventry, et al_., 1935),
The 7-day LC5o for rainbow trout was 0.08 mg/1 with an estimated median
period of survival of one year at 0.004 mg/1 (Merkens, 1958). Rainbow
trout were shown to avoid a concentration of 0.001 mg/1 (Sprague and
Drury, 1969). Dandy (1972) demonstrated that brook trout had a mean
survival time of 9 hours at 0.35 mg/1, 18 hours at 0.08 mg/1 and 48
hours at 0.04 mg/1, with mortality of 67 percent after 4 days at 0.01
mg/1. Pike (1971) observed a 50 percent brown trout mortality at
0.02 mg/1 within 10.5 hours and 0.01 mg/1 with 43.5 hours.
The range of acutely lethal residual chlorine concentrations is
narrow for various species of warm Water fish. Arthur (1972) determined
-------
96-hour LCgQ values for the walleye, black bullhead, white sucker,
yellow perch, largemouth bass, and the fathead minnow. The observed
concentration range was 0.09 to 0.30 mg/1.
Using fathead minnows in a continuous bioassay technique at
Michigan treatment plants, Zillich (1972) found that an average concen-
tration of 0.16 to 0.21 mg/1 killed all of the test fish and that
concentrations as low as 0.07 mg/1 caused some mortalities. Pyle (1960)
demonstrated a 50 percent mortality of smallmouth bass exposed to 0.5
mg/1 within 15 hours. The mean 96-hour LC5g value for golden shiners
was 0.19 mg/1 (Esvelt, ejt al_., 1971). Arthur and Eaton (1971), working
with fathead minnows and the freshwater crustacean, Gannnarus pseudolimnaeuSj
in dilute wastewater, found that the 96-hour LCgQ of total residual
chlorine for Gammarus was 0.22 mg/1 and that all fathead minnows were
dead after 72 hours at 0.15 mg/1. At concentrations of .09 mg/1, all
fish survived until the seventh day when the first death occurred. In
exposure to 0.05 mg/1 residual chlorine, these investigators found
reduced survival of Gammarus and at 0.0034 mg/1 there was reduced
reproduction. Growth and survival of fathead minnows after 21 weeks
were not affected by continuous exposure to 0.043 mg/1 residual chlorine.
The highest level showing no significant effect was 0.016 mg/1. Working
with secondary wastewater effluent, Arthur (1972) found that repro-
duction by Gammarus was reduced by residual concentrations above 0.012
mg/1 residual chlorine.
In marine water, 0.05 mg/1 was the critical chlorine level for
young Pacific salmon exposed for 23 days (Holland, £t al_., 1960). The
tff
-------
lethal threshold for Chinook salmon and coho salmon for a 72-hour
exposure was noted by these investigators to be less than 0.01 mg/1
chlorine. Studies on the effect of residual chlorine to marine
phytoplankton indicate that continuous exposure to 0.10 mg/1 reduced
primary production by 70 percent while 0.2 mg/1 for 1.5 hours reduced
primary production by 25 percent (Carpenter, e£ a_l_., 1972). Laboratory
studies on ten species of marine phytoplankton indicate that 8 50
percent reduction'in growth rate occurred at chlorine concentrations of
0.075 to 0.250 mg/1 during a 24-hour exposure period (Gentile, ejt aj_.,
1973). Oysters are sensitive to chlorine concentrations of 0.01 to
0.05 mg/1 and react by reducing pumping activity. At chlorine concen-
trations of 1.0 mg/1, effective pumping could not be maintained (Galtsoff,
1946).
Chlorine residuals of 10 ug/1 have been shown to kill adult
salmonid fish in a period of several days in fresh water and the fry
of these species have been killed in chlorine residuals of 6 ug/1.
The criterion of 2 ug/1 chlorine should afford protection to this group
of fish when exposed on a continuing basis. Consi4ering the data
presented above, a criterion of 10 ug/1 should afford protection to
other freshwater fish and marine aquatic life (Brungs, In Press). Brungs
(1973) reported that aquatic organisms may tolerate short-term exposure
to higher levels of residual chlorine than the concentrations which have
adverse chronic effects. Basch and Tfcichan (In Press) have shown that
repeated daily exposure at these levels will have toxic effects on aquatic
life.
US'
-------
REFERENCES CITED:
Arthur, J.W., 1972. Progress Report, Natl. Water Quality Lab.,
Environmental Protection Agency, Duluth, Minn.
Arthur, J.W., and L.G. Eaton, 1971. Chloramine toxicity to the amphipod,
Gammarus pseudolimnaeus, and the fathead minnow, Pimephales promelas.
Jour. Fish. Res. Bd. Can., 28:1841.
Basch, R.E. and J.G. Truchan, In Press. Tenacity of chlorinated
power plant condenser cooling waters to fish. U.S. Environmental
Protection Agency, Ecological Research Series. Grant R800700
Michigan Water Resources.
Brungs, W.A., 1973. Effects of residual chlorine on aquatic life.
Jour. Water Poll. Cont. Fed., 45:2180.
Brungs, W.A., In Press. Effects of wastewater and cooling
water chlorination on aquatic life. U.S..Environmental
Protection Agency, Ecological Research Series.
-------
Carpenter, E.J., et a]_., 1972. Cooling water chlorinatlon and
productivity of entrained phytoplankton. Marine Biology, 16:37.
Coventry, F.L., e_t a_K, 1935. The conditioning of a chloranrfne treated
water supply for biological purposes. Ecology, 16: 60.
Dandy, J.W.I., 1972. Activity response to chlorine in brook trout,
Salvelinus fontinalis. Can. Jour. Zool., 50:405.
Esvelt, L.A., ejt aJL, 1971. Toxicity removal from municipal wastewaters,
Vol. IV, A study of toxicity and biostimulation in San Francisco
Bay-Delta Water. SERL Rept. No. 71-7, San Eng. Res. Lab., Univ. of
California, Berkeley.
Galtsoff, P.S., 1946. Reactions of oysters to chlorination. Fish
and Wildlife Service, USDI Research Report 11.
-------
Gentile, J.H., e_t al_., 1973. The effects of chlorine on the growth
and survival of selected species of estuarine phytoplankton and
zooplankton. 36th Ann. Meeting Amer. Soc. Limnol. and Oceanog.
Holland, G.A., et al_., 1960. Toxic effects of organic and inorganic
pollutants on young salmon and trout. State of Wash., Dept. of
Fisheries, Res. Bull. No. 5:198.
Merkens, J.C., 1958. Studies on the toxicity of chlorine and chloramines
to the rainbow trout. Water Waste Trt. Jour., (C.B.), 7:150.
Michigan Department of Natural Resources, 1971. Chlorinated municipal
waste toxicities to rainbow trout and fathead minnows. Water Poll.
Cont. Res. Ser., 18050 GZZ, EPA, Washington, D. C.
Moore, E.W., 1951. Fundamentals of chlorination of sewage and waste.
Water and Sewage Works, 98:130.
Pike, D.J., 1971. Toxicity of chlorine to brown trout. New Zealand
Wildlife, No. 33.
Pyle, E.A., 1960. Neutralizing chlorine in city water for use in fish
distribution tanks. Progressive Fish Cult., 22:30.
Sprague, J.B. and D.E. Drury, 1969. Avoidance reactions of salmonid
fish to representative pollutants. In: Advances in water pollution
research. Proc. 4th Intl. Conf. Water Poll. Res., Pergamon Press,
London* England.
-------
Tsai, C., 1973. Water quality and fish life below sewage outfalls.*
Trans. Amer. Fish. Soc., 102:281.
Zillich, V.A., 1972. Toxicity of combined chlorine residuals to
freshwater fish. Jour. Water Poll. Cont. Fed., 44:212.
fed
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CHROMIUM
CRITERIA:
50 ug/1 for domestic water supply (health);
100 ug/1 for freshwater aquatic life
INTRODUCTION:
Chromium Is the seventeenth most abundant nongaseous element in the
earth's crust (Schroeder, 1970); Its concentration range In the continental
crust Is 80 to 200 mgAg/ with an average of 125 mgAg (N&S/ 1974a).
Although chromium has oxidation states ranging from Cr~^ to Cr+^, the
trivalent form most commonly Is found in nature. Chromium is found
rarely in natural waters, ranking twenty-seventh or lower among the
elements in seawater and generally is well below 1 ug/1. Kopp (1969)
reported that for 1,577 surface water samples collected at 130 sampling
points in the United States, 386 samples contained from 1 to 112 ug/1;
the mean was 9.7 ug/1 chrctnium. Durum, et al.,(1971) in a similar survey
of 700 samples, found that none contained over 50 ug/1 of hexavalent
chromium and 11 contained over 5 ug/1. Chromium is found in air, soil,
some foods, and most biological systems; it is recognized as an essential
trace element for humans (NAS, 1974a).
RATIONALE:
The earliest consequences of mild chromium deficiency in experimental
animals is a reduced sensitivity of peripheral tissues to insulin; more
severe deficiency in rats and mice results in fasting hyperglycemia,
-------
glycosuria, and mild growth retardation that 1s probably due, at least
In part, to reduced insulin activation. Glucose intolerance is a common
human problem, and one of its many possible causes is chromium deficiency
(NAS, 1974a).
The biological activity of chromium, i.e., its effect as an essential
metal, is restricted to its trivalent state, Cr3+. The Cr3+ ion forms
complexes that are stable at or below pH 4, but which readily hydrolyze
at "high" pH values; resulting in the formation of polynucleate bridge
qj.
complexes. At the normal pH of blood, Cr exists in large, insoluble
macro-molecules which precipitate and become biologically inert; Cr3+
must therefore be supplied as a complex of suitable stability in order
to be utilized.
Because of the present inadequacy of knowledge about the forms and
biological availability of chromium in foods, it is not possible to
quantify human dietary requirements. It is known that adult urinary
loss is 5 to 10 ug/day, and since this constitutes the major portion of
the daily loss, at least this much must be replaced to maintain balance.
Based on rat studies indicating that absorption of various Cr3+ compounds
can range from below 1 percent to 25 percent of a given dose, and assuming
that the human case is similar, a dietary intake of 20 to 500 ug/day
would balance urinary losses. It is estimated that daily chromium
intakes in the U.S. are marginal and vary from 5 to 100 ug (WHO, 1973);
it seems unlikely that any Cr3+ ingested via public drinking water would
be appreciably assimilated.
-------
No harmful effects were observed when food or water containing
moderate amounts of Cr^+ was administered to laboratory animals, e.g.,
cats that were fed chromic phosphate or oxydicarbonate at 50 to 1000
mg/day for 80 days, or rats drinking water containing 25 mg/1 for a year
or 5 mg/1 throughout their lives (NAS, 1974a).
Hexavalent chromium, on the other hand, is irritating and corrosive
to the mucous membranes; it is absorbed via ingestion, through the skin,
and by inhalation, and is toxic when introduced into laboratory animals
systemically (NAS, 1974a). Knowledge of the harmful human health
-effects of hexavalent chromium has been obtained almost entirely from
occupational health effects. Lung cancer, ulceration and perforation of
the nasal septum, and a variety of other respiratory complications and
skin effects have been observed.
Symptoms of excessive dietary intake of chromium in man are unknown,
and chromium deficiency is of greater nutritional concern than over-
exposure. Chromium is the only element whose tissue concentration
appears to decline with increasing age in the U.S. population; lung
concentrations do not decline with age, suggesting that lung chromium is
not in equilibrium with that in the rest of the body (NAS, 1974a).
Berg and Burbank (1972) compared concentrations of eight carcino-
genic trace metals in water supplies (after Kopp and Kroner, 1967) with
State cancer mortalities for major U.S. water basins, and no significant
correlations were found for chromium. It should be emphasized that
these data are not conclusive.
•71
-------
The U.S. Public Health Service Drinking Water Standard. (USPHS,
1962) states that the presence of hexavalent chromium in excess of 0.05
mg/1 shall constittrte grounds for rejection of the supply', and no harmful
human health affects have been reported at this level. The NAS/NAE
Committee on Water Quality Criteria recommended (NAS, 1974) that public
water supply sources for drinking water contain no more than 0.05 mg/1
total chromium, largely on the basis that lifetime tolerable levels of
chromate ion are not known for man. There are insufficient data on the
effect of the defined treatment process on chromate chromium removal.
Chromium as Cr3+ is not likely to be present in waters of pH 5 and above
because the hydrated oxide is very sparingly soluble.
A family of four individuals is known to have drunk water for
a period of 3 years at a chromium level of 0.45 mg/1 without known
effects on their health as determined by a single medical examination
(Davids and Lieber, 1951). A study was designed by MacKenzie, e_t £]_.
(1958) to determine the toxicity of hexavalent and trivalent chromium
ions to rats at various drinking water levels. After one year at levels
of 0.45 to 25 mg/1, this study showed no evidence of toxic response
In body weight, food consuntption, blood changes, or mortality. However,
significant accumulation of chromium occurred in the tissues at con-
centrations greater than 5 mg/1. Recent studies (Naumova, 1965)
demonstrated that 0.1 mg of K,gCr207/kg enhances the secretory and motor
activity of the dog intestine.
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From these and other studies (Gross and Heller, 1946; Brard, 1935;
Conn, ejt aJL, 1932; Schroeder, ejt al_., 1963; Schroeder, et al_., 1963a),
it appears that a concentration of 50 ug/1 of chromium in domestic water
supply incorporates a reasonable safety factor to avoid any hazard to
human health.
In addition, the possibility of dermal effects from bathing in
water containing 50 ug/1 chromium would likewise appear remote, although
chromium is recognized as a potent skin sensitizer. Domestic water
supplies should, therefore, contain no more than 50 ug/1 total chromium.
Fish appear to be relatively tolerant of chromium, but some aquatic
invertebrates are quite sensitive. Toxicity varies with species, chromium
oxidation state, and pH. Pickering and Henderson (1966) conducted
static bioassays with four warm water fish species; fathead minnows,
Pimephales promelas; bluegill, Lepomis macrochirus; goldfish, Carassius
auratus; and guppies, Poecilia reticulata. They obtained soft water
(hardness - 20 mg/1; alkalinity - 18 mg/1; pH - 7.5) 96-hour LC_n values
bO
for hexavalent chromium ranging from 17.6 mg/1 for fathead minnows to
118 mg/1 for bluegill; hard water (hardness - 360 mg/1; alkalinity -
300 mg/1; pH - 8.2) 96-hour LC values for hexavalent chromium ranging
ou
from 27.3 rag/1 for fathead minnows to 133 mg/1 for bluegill. Their 96-hour
LCg trivalent chromium values (chromium potassium sulfate) ranged from
3.33 mg/1 for guppies to 7.46 mg/1 for bluegill in soft water. The
LC50 for fathead minnows exposed to potassium chromate in soft water
was 45.6 mg/1.
73
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Pickering (NAS, 1974) found 96-hour iJCLv. and safe hexavalent
chronium concentrations of 33 mg/1 and 1.0 mg/l, respectively, for fathead
minnows in hard water. Pickering's Cr data for fathead minnows in
hard water showed a 96-hour I£50 of 27 mg/1 and a safe cjonoentration of
1.0 mg/1. Benoit (In Presfc) obtained soft water (45 mg/1 hardness) 96-
hour IC50 and safe hexavalent chromium values of 59 mg/1 and 0.2 mg/1,
respectively, for brook trout, Salvelinus fontinalis, and 69 amd 0.2 mg/1/
respectively for rainbow trout, Salmo gairdneri.
Olson (1958) found the growth and survival of chinook salmon,
Oncorhynchus kisutch, alevins and juveniles to be significantly reduced
at hexavalBBt chromium concentrations of 0.2 mg/1. He saw no detrimental
effects on salmon alevins at trivalent chromium concentrations of 0.2 mg/1.
Beisinger and Christensen (1972) reported a 16 percent reproductive
impairment in Daphnia roagna in soft water (45 mg/1 as CaOO ) at a
concentration of 0.33 mg/1 of trivalent chromium.
The data available indicate hexavalent chromium to be somewhat more
toxic than trivalent chromium in the case of chinook salmon/ and since
significant effects were seen on fish at 0.2 rag/1 of hexavalent chromium
a recommended criterion of 0.10 mg/1 should provide adequate protection
for both freshwater invertebrates and fish.
Lowman, et al. (1971) reported marine chrcmium concentration factors
of 1,600 in benthic algae, 2,300 in phytoplankton, 1,900 in zooplankton,
440 in molluscan soft parts, and 70 in fish muscle.
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Raymont and Shields (1964) reported chromium threshold toxicity
levels of 5 mg/1 for small prawns, Leander squilla, 20 mg/1 (as
for the shore crab, Carcinas maenus, and 1 mg/1 for the polychaete,
Nereis virens. Doudoroff and Katz (1953) found that mummichogs,
Fundulus heteroclitus. tolerated 200 mg/1 K2Crg04 in seawater for over
a week.
Holland, e^aJL (1960) reported that 31.8 mg/1 chromium (as K^CKfy)
in sea water was 100 percent fatal to coho salmon, Oncorhynchus kisutch,
and Gooding (1954) reported that 17.8 mg/1 hexavalent chromium in sea-
water was toxic to the same species.
Clendenning and North (1960) showed that 5.0 mg/1 hexavalent chromium
reduced photosynthesis by 50 percent in the giant kelp, Macrocystis
pyrifera, during 4 days of exposure.
Based on the foregoing discussion of the marine toxicity of chromium,
its accumulation at all trophic levels, and the sensitivity of lower
aquatic forms to chromium, the freshwater criterion should protect
marine populations.
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REFERENCES CITED
Benoit, D.A., 1976. Toxic effects of hexavalent chromium on brook
trout (Salvelinus fontinalis) and rainbow trout (Saliro gairdneri)
Water Research, (In Press).
Berg, J.W. and F. Burbank, 1972. Correlations between carcinogenic tre,ce
metals in water supplies and cancer mortality. Ann. N.Y. Acad. Sci.,
199: 249.
Biesinger, K.E. and G.M. Christensen, 1972. Effects of various metals on
survival, growth, reproduction and metabolism of Daphnia magna.
Jour. Fish Res. Bd. of Canada, 29: 1691.
Brard, D., 1935. Toxicological study of certain chromium derivatives.
II. J. Pharm. Chem., 21: 5.
Clendenning, K.A. and W.J. North, 1960. Effect of wastes on the giant kelp,
Hacrocystis pyrifera. Proc. 1st Int'l. Conf. on Waste Disposal in the
Marine Environment. Pergamon Press, New York, p. 82.
Conn, L.W., eit al_., 1932. Chromium toxicology. Absorption of chromium by the
rat when milk containing chromium was fed. Amer. Jour. Hyg., 15:760.
Davids, H.W. and K. Lieber, 1951. Underground water contamination by
chromium wastes. Water and Sewage Works, 98: 528.
Doudoroff, P. and M. Katz, 1953. Critical review of literature on the toxidty
of industrial wastes and their components on fish. The Metals as Salts.
Sew. and Ind. Wastes, 25:802.
-------
Durum, W.H., £t al_., 1971. Reconnaissance of selected minor elements in
surface waters of the United States, October 1970. U.S. Geological Survey
Circular 643. Wash., D.C. U.S. Geological Survey.
Gooding, D., 1954. Pollution research, toxicity studies. Sixty-fourth Annual
Report, Washington State Department of Fish. Olympia, Washington.
Gross, W.G. and V.G. Heller, 1946. Chromates in animal nutrition.
Jour. Ind. Hyg. Toxicol., 28:52.
Holland, G.A., jrt al_., I960. Toxic effects of organic and inorganic
pollutants on young salmon and trout. Washington Department of Fish
Res. Bull. No. 5. Seattle, Washington.
Kopp, J.F., 1969. The occurrence of trace elements in water, In: D.D. Hemphill,
Ed., Trace Substances in Environmental Health - III. Proc. Univ. of
.Missouri's 3rd Annual Conference on Trace Substances in Environmental
Health. Univ. of Missouri, Columbia, p. 59.
Kopp, J.F. and R.C. Kroner, 1967. Trace metals in waters of the United
States. U.S. Department of the Interior, Federal Water Pollution Control
Administration, Cincinnati, Ohio.
Lowman, F.G., et_ al_., 1971. Accumulation artd redistribution of radionuclides
by marine organisms, In: Radioactivity in the Marine Environment.
National Academy of Sciences, Wash., D.C. p. 161.
MacKenzie, R.D., et^ a_l_., 1958. Chronic toxicity studies. II: Hexavalent
and trivalent chromium administered in drinking water to rats. A.M.A.
Archives of Industrial Health, 18: 232.
National Academy of Sciences, National Academy of Engineering/ 1974. Water
quality criteria, 1972. U.S. Government Printing Office, Washington,
D.C. 77
-------
National Academy of Sciences, 1974a. Chromium. U.S. Government Printing
Office, Washington, D.C.
Naumova, M.K., 1965. Changes in the secretory and motor functions of the
intestine following administration of potassium dichromate..
Gig. Tr. Prof. Zabol., 9: 52.
Olson, P.A., 1958. Comparative toxicity of Cr(VI) and Cr(lll) in
salmon. Hanford Biology Research Annual Report for 1957,
January 10, 1958. HW-53500, p. 215.
Pickering, Q.H., and C. Henderson, 1966. The acute toxicity of some heavy
metals to different species of warm water fishes. Int'l. Jour.
Air-Water Pollution, 10: 453.
Raymont, J.E.G., and J. Shields, 1964. Toxicity of copper and chromium 1n
the marine environment, In: E.A. Pearson, Ed., Advances in Water
Pollution Research, Proc. 1st Int'l. Conf. Macmillan Co., New York,
Vol. 3, p. 275.
Schroeder, H.A,., et^ a]_., 1963. Effects of chromium, cadmium and other trace
metals on the growth and survival of mice. Jour. Nutrition, 80:39.
Schroeder, H.A., £t al_., 1963a. Effects of chromium, cadmium and lead on the
growth and survival of rats. Jour. Nutrition, 80:48.
Schroeder, H.A., 1970. Chromium. Air Quality Monograph #70-15. American
Petroleum Institute, Wash., D.C.
USPHS, 1962. Public Health Service Drinking Water Standards (rev. 1962).
PHS Publication 956. Wash., D.C.
MHO, 1973. Chromium, In: Trace Elements in Human Nutrition. World Health
Organization Technical Report Series No. 532. Geneva, p. 20.
78
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FECAL COLIFORM BACTERIA
CRITERION!
Bathing Waters
Based on a minimum of not less than five samples taken over
a 30-day period, the fecal ooliform bacterial level should not exceed
a log mean of 200 per 100 ml, nor should more than 10 percent of the
total samples taken during any 30 day period exceed 400 per 100 ml.
Shellfish Harvesting Waters
Not to exceed a median fecal collform bacterial concentration
of 14 MPN per 100 ml with not more than 10 percent of
samples exceeding 43 MPN per 100 ml for the taking of
shellfish.
INTRODUCTION'.
It was recognized even before the mlcroblal etiology of
disease was known, that water can serve as a medium for the
transfer of disease. The cause-effect relationship of disease
transmission and specific fecally-assodated microbes was
initially defined by Von Fritsch 1n 1880, when he Identified
Klebsiella spp. 1n human feces. Further confirmation of the
79
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relationship between fecally-assodated microbes and potential
disease was developed by Escherich when he descried .Bacillus .mlj
(Escherlchla col 1) as an Indicator of pollution (Wolf,, 1972;
Guarrala, 1972). Since these early observations, the role of
biological Indicator organisms 1n defining water quality has
become essential and addresses three categories generally: to
Identify environmental changes; to quantify pollution levels; or
to be used 1n laboratories to study under controlled conditions
phenomena which could be extrapolated to the environment (Butler,
e_t al_., 1972).
Microbiological Indicators have been used to determine or
Indicate the safety of water for drinking, swimming and shellfish
harvesting. As our knowledge concerning microbiology has
Increased, so has our understanding of the complex
Interrelationship of the various organisms with disease. Viruses
causing a number of diseases and non-fecally associated bacteria
causing Infections of the ear, eye, nose and throat all have been
Isolated from water (Bonde, 1974^ Scarplno, 1974). The
relationship between numbers o'f specific disease* causing
organisms 1n water and the potential for transmission of disease
remains elusive since the number of organl sms/r,equ1 red to cause
disease varies depending upon the organism, the host, and
the manner 1n which the bacteria and host Interact. For example,
1n some Instances a single cell of Salmonella, or a single plaque-
forming unit (PFU) may be all that 1s necessary to cause a
80
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disease; however, in other instances the nimbers of bacteria
necessary to cause an illness may be up to 10 to 107 or even more
viable organisms. The use to which water is put (i.e., swimming
or drinking) the type of water, (marine or fresh), and the geographical
location are all factors to be weighed in determining safe micro-
biological criteria.
Ideally, a microbiological Indicator organism should fulfill
all of the following criteria: (1) It should be applicable to all
types of water, *(2) It should be present whenever pathogens are
present with a survival time equal to that of the hardiest
enteric pathogen, and (3) the Indicator should not reproduce 1n
contaminated waters thus resulting 1n Inflated values (Scarplno,
1974). Unfortunately, no such Indicator organisms are known.
Use has been made of coliform or fecal coliform bacteria as indicators
of pollution.
Bacteria of the coliform group are considered the primary
Indicators of fecal contamination and are one of the most
frequently applied Indicators of water quality. The coliform
group 1s made up of a numb'er of bacteria Including the genera
Klebslella, Escher1ch1a, Serratla. Erwlnla and Enterobacterla.
Total coliform bacteria are all gram negative asporogenous rods
and have been associated with feces of warmblooded animals.
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and with soil. The fecal coliform bacteria, which comprise
a portion of the total coliform group, are able to grow
at 44.5° C and ferment lactose producing acid and gas. Use
of fecal coliform bacteria has proven to be of more sanitary
significance than the use of total coliform bacteria because
they are restricted to the intestinal tract of warm-blooded
animals and are now used to define water quality for swimming.
Arguments have been advanced for the use of Escherichia coli
as the indicator of choice for fresh fecal pollution (Bonde,
1966). However, the methods for its identification require
time, complex biochemical reactions, and experienced micro-
biologists, and the information gained by the additional work
and expense involved in its identification may not provide
sufficient additional information above the elevated temperature
test (for fecal coliform bacteria) to warrant its use.
Enterococci were recognized as early as 1890 as being indicators
of recent fecal pollution from warm-blooded animals (Geldreich and
Kenner, 1969). Hie enterococci possess many characteristics which make
them an ideal indicator system since they do not multiply in
the aquatic environment and are serologically characteristic.
However, there are numerous biotypes, and no good standard-
ized method is available for biochemical testing. Such
methodology is essential since identification is based primarily
upon biochemical characterization.
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Additional tn1croh1al systems have also been proposed as
potential Indicators of fecal pollution. Clostridium
perfringens, an anaerobic spore former, has been associated
frequently with sewage pollution. However, 1n a suspension of
fecal material the ratio of Cl. perfrlngens to E_. coll may differ
from that 1n the effluent from a sewage treatment plant, 1n
seawater or In sludge. Also the spores of Cl. perfrlngens are
resistant and will vegetate upon culturing, hence 1t would be
difficult to associate their presence with recent pollution.
Presence of microorganisms other than fecal collform bacteria may
also be indicative of water quality; however, the strict
correlation between a pollutant problem other than fecal
pollution and the numbers of the indicator organism 1s not always
clear. The pseudomonads that are comprised of numerous free
living saprophytes found in both fresh and marine waters have
only one of Its members as a known pathogen^ Pseudomonas
^eruginosa, which may be an Indicator of pollution from the
presence of warm-blooded animals jRingem and Drake, 1952; Taylor,
1968; Reltler and Seligaman, 1957). While the number of P..
aeruqinosa in sewage usually 1s quite low and occurs too
sporadically to be of value as an indicator organism for fecal
pollution (Bonde, 1974), this organism may be an Indicator of
human pollution other than fecal.
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Bacterial pathogens, which occur in bathing waters
and may cause disease even when not Ingested are Klebslella
pneumonlate and Pseudomonas aeruginosa.
RATIONALE:
Bathing Waters
Pollution of aquatic systems by the excreta of warnblooded
animals creates public health problems for man and animals and
potential disease problems for aquatic life. It 1s known that
enteric mlcroblal pathogens may Inhabit the gut of most
warmblooded animals and are shed in feces. The presence of
bacterial, viral, protozoan, and possibly funqal species which
are either pathogens for or possess the potential to Infect man
and other organisms 1s Indicated hy the presence of the fecal
collform group of bacteria. Thus, the number of fecal conforms
present 1s Indicative of the deqre'e of health risk associated
with using the water for drlrtklng, swimming, or shellfish
harvesting.
Arguments against the use of fecal collform bacteria to
define swimming quality 1n waters have noted the paucity of
ep1dem1olog1cal evidence Unking fecal collform levels 1n bathing
waters and the Incidence of disease (Moore, 1959; 1971).
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A problem of potential medical significance 1s the transfer
of characteristics which will alter the resistance of pathogenic
bacteria, I.e. the R-factor (RTF - resistance transfer factor),
to certain antibiotics, heavy metals and ultraviolet light. The
significance of this, while not fully known, suggests that human
pathogens1n water may become resistant to common antibiotic
therapy once the bacteria infect man or animals. One example of
this problem has been already reported. Recently an outbreak of
typhoid fever was found to be caused by Salmonella typhi
containing the chloramphenical and amphicillin resistant factor
(Datta and Olearte, 1974; Olearte and Galindo, 1973).
Disease transmission via the aquatic route including drinking
water, recreational water, and seafood from polluted water, has been
and continues to be a problem. Presently, the indicator systems
considered to be the most practical are the coliform and the fecal
coliform groups. Additional microbiological problems, which can be
anticipated by the presence of specific fcacteria or viruses,
are recognized. Correlation between human pollution sources
and the numbers and significance of the microbial system in question remain elusive.
Nevertheless, as the relationships become clear additional criteria for water
quality will evolve. Berg (1974) has shown that Polio Virus ' I at a concentration
of 2 plaque forming units (PFU), a standard means for measuring virus concentrations
in tissue culture, will cause disease in 67 percent of the uninncculated population.
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However, the lack of epidemic!oglcal correlation between fecal
conform levels in coastal swimming waters and the Incidence of
disease may not have validity In fresh waters and it does not
take Into account non-reported diseases which may
develop as an unrecognized result of swimming 1n polluted waters.
Ep1dem1olog1cal evidence 1s but one consideration 1n setting
microbiological criteria. The presence of fecal collform
bacteria Indicates degradation of water quality and a relative
risk of disease transmission.
In studies conducted in Lake Michigan at Chicago, Illinois
(Smith, e_t al_., 1951), 1n an Inland, river (Smith and Woolsey,
1952) and in tidal waters at Long Island, New York (Smith and
Woolsey, 1961), a statistically significant Increase 1n the
Incidence of illness was observed among swimmers who used the
Lake Michigan beaches on selected da^s and the Ohio River beach
of poorest water quality. The mean total collform bacteria
content of fresh waters was 2,300 per 100 ml and 2,700 per 100 ml,
respectively. No relationship between the total collform levels
and swimming-related diseases was found at the ocean beach.
These studies demonstrated that an appreciably higher overall
Illness Incidence may be expected among ^winners when compared to
non-swimmers, but the data are inconclusive. The data provide a
positive correlation between total collform numbers and the
Increased risk of disease associated with swimming 1n these
waters. The diseases found were Infections of the eye, ear.
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nose, and throat, as well as Intestinal ailments (Stevenson,
1953).
Outbreaks of typhoid fever (Salmonella typhl) have been
associated with swimmers 1n "heavily" polluted beaches 1n western
Australia (Kovacs, 1959; Snow, 1959). In this case, a sewer
outfall pipe was located one mile from the hathlnq beach where
the typhoid outbreaks occurred and the sewer was overloaded
because of rapid population growth. Use of a log mean of 200
fecal collform bacteria per 100 ml, with the provision that 10
percent of the total samples during a thirty-day period not
exceed 400 fecal conforms per 100 ml, allows for variations 1n
environmental conditions such as shifts 1n wind direction,
current flow, and tidal fluctuations. At levels above the 200
fecal conforms per 100 ml the risk of exposure to pathogenic
microbes Increases. Correlation between the fecal collform level
and microblal pathogen levels 1s the Important relationship,
since the fecal conforms themselves serve as an Indicator of the
quality of water 1n relation to_ fecally associated microblal
pathogens. Direct demonstration of the microblal pathogens Is
not always feasible.
Detection of Salmonella has shown that 1n fecally polluted
marine waters the level may vary between 1 and 1,000 per liter
(McCoy, 1964). Occurrences of viruses 1n ocean waters at levels
of 60 plaque-forming units (PFtl) per liter have been found
87
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(Shuval, £t,aU» 1971). It 1s estimated that these values may
represent only 10 to 50 percent of the total viral pathogens
present because of the limited recovery efficiency of the
methods used. Also, other viruses associated with feces may be
present but may not grow in the tissue culture systems used.
Another Important consideration 1n determining the safe
microbiological criterion 1s the minimum dosage necessary to
Infect a bather. As few as 3 to 5 organisms of S_. typhosa have
been reported to cause Infection, whereas, 1 x 105 to 1x10 cells
for other Salmonella serotypes may .be required to produce
disease. Similarly, massive concentrations of cells of
enteropathogenlc E_._ coli have been reported necessary to produce
Infections in adult volunteers. Alteration of gastric function
either by raising the pH 1n the stomach or by facilitating
gastric emptying can reduce this dosage several orders of magnitude.
Recent data from experiments with adult volunteers indicate that the
infectious dose for Shlgella flexnerj 2a 1s less than 200 cells
(Geldrelch, 1974). Although enteric viruses are found 1n
relatively small numbers 1n polluted waters, their occurrence
could be hazardous since the minimum Infective dtose for humans
has not been firmly established. For children, the minimum
Infective dose may be 1 or 2 PFU (Plotkln and Katz, 1967), while
for adults the minimum infective dose may be the same or higher.
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However, less is known concerning the minimal infective dose for
pathogenic bacteria that cause, ear, eye, nose, and throat ailments
(Geldreich, W74).
Use of fecal collform bacteria as a single parameter for
monitoring recreational water quality must ultimately relate to
the probable occurrence of waterborne pathogens. Currently, the
only relationship which has been developed between the fecal
collform Indicator and waterborne pathogens is that of
Salmonella to fecal collform density Data which have been
developed Indicate a sharp Increase 1n the frequency of
Salmonella detection when fecal collform densities are above 200
organisms per 100 ml of freshwater. When there are over 200
fecal conforms per 100 ml, Salmonella Isolation should approach
100 percent frequency. Data from estuarlne waters were grouped
to Include the level of 1 to 70 fex:al conforms that 1s of
Interest to Investigators of^shellfish waters. For this range,
below 70 fecal cOllforms per 100 ml, only 6.7 percent of the 184
/
Salmonella examinations w-ere positive. At the 200 fecal collform
level, the 28.4 percent occurrence of Salmonella 1n estuarlne
» ""*"""™""~~—
waters was essentially Identical with data compiled from
freshwater environments. In polluted estuarlne waters containing
fecal collforms ranging from 201 to above 2,000, a recovery of
89
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Salmonella was seen 60 percent of the time. The recovery rate of
Salmonella 1n estuarine waters 1s lower than that in fresh waters.
In fresh water, Salmonella were recovered 85 to 98 percent of the
time, 1n the range of from 201 to 2,000 fecal conforms. The
lower value for Isolation of Salmonella from estuarine waters may
be related to limitations of Salmonella methodology (Geldrelch,
1972). It must be noted that any projection of the qualitative
recoveries of Salmonella Into a comparison with fecal collform
quantification has recognized limitations. Such projections do
nevertheless suggest that the 200 fecal collform per ion ml limit
for recreational waters is a useful water quality value.
Evaluation of the microbiological suitability for marine and
fresh waters should be based upon the fecal coliform
levels. As determined by either the multiple-tube fermentation
for marine water or the membrane filter method for fresh water,
and based upon a minimum of not less than five samples taken
over not more than a 30-day period, the fecal coliform bacterial
level should not exceed a log mean of 200 per 100 ml, nor
should more than 10 percent of total samples during any 30-day
period exceed^ 400 per 100 ml.
Shellfish
Shellfish, being filter feeders, require a high quality of
water 1n order to be microblologically safe for human consumption
90
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as either raw or partially cooked. Fecal collform bacteria,
other bacterial pathogens, and viruses found 1n water and
sediments are concentrated by shellfish, depending upon
temperature, density of pathoqens, currents, depth, water
chemistry, and shellfish feeding activity (Van Donsel and
Geldrelch, 1971; Metcalf and Stiles, 1968). Once concentration
of pathogens occurs, flushing of microorganisms will not
necessarily occur at the same rate (Janssen, 1974; Kelly and
Arcisy, 1954). Because of the established relationship between
coliform levels and enteric pathogens, shellfish waters
historically have been classified on the basis of total coliform
levels.
An attempt by the National Shellfish Program'has been made to
correlate fecal collform bacteria to enteric pathogens. Ideally,
any specific fecal collform bacterial limit for shellfish should
be based on a correlation with pathogenic occurrences 1n the
aquatic environment and with en1dem1olog1cal evidence of
Increased health risk among shellfish consumers. However, these
data are not available and calculations have been based on health
risks Incurred through Increased pathogen occurrence 1n waters of
differing levels of fecal contamination. Recent data In
shellfish growing waters have shown that Salmonella occurred 1n
4.7 percent of water samples having fecal collform densities of 1
to 29 per 100 ml (Slanetz e_t aj_., 1974). Oysters growing In
these waters accumulated from 33 to 2200 fecal collform bacteria
-------
per 100 grams of shellfish meat, with Salmonella occurrence at
6.1 percent.
Shellfish contamination ±s intensified further by the normal
accumulation of waterborne organisms in bottom sediments through
the action of sedimentation. Investigation of both the overlylnq
water and bottom sediments from lakes and streams-has indicated
a 100- to 1000-fold Increase 1n fecal collform densities at the
water-sediment Interface (Van Donsel and Geldrelch, 1971).
Enteric viruses (Coxsackle B3) 1n the bottom sediment of
shellfish growing waters along the New Hampshire estuary have
been found when the fecal conform densities were as low as 10
organisms per 100 ml 1n the overlying waters (Slanetz ejt al_., 1965)
Indicators of fecal pollution more specific, than the total
conforms 1n shellfish waters have been sought. Candidate
organisms or groups of organisms Include E_. coll and the fecal
collform group. The fecal conforms have a higher positive
correlation with fecal contamination from all warm blooded
animals than does £. col 1 (Geldrelc'h, 1974). Usually £. cpJH. 1s
the'most numerous bacterium of the fecal collform group; however,
under some conditions other fecal conforms may predominate
(Sears e_t aj_., 1950). Analysis of data comparing the correlation
of fecal conforms ~* and £• £2.11 to fecal
collform sources confirms the fact that the fecal coliforms
-------
reflect sanitary quality of water. In comparing results of the fecal
COl 1form test to those of the E. ooli procedure in shellfish waters, E.
col 1 was reported to range from 75 to 90 percent of the fecal
collform population with the fecal conform bacteria giving a
96.6 percent correlation (Presnell, 1974). Therefore, the use of
the fecal collform test avoids the undesirable risk of excluding
some fecal contamination.
The microblolog i.cal criterion for shellfish water quality has
been accepted by international agreement to be 70 total collforms
per 100 ml, using a median MPN, with no more than 10 percent of
the values exceeding 230 total coliforms. No evidence of
epidemiological 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 collform occurrences
can be associated with fecal poMutlon (Gallagher, et_ aj_., 1969).
Thus, attention has been directed toward the adoption of the
fecal collform test to measure nore precisely 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
93
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conforms to numbers of fecal collforms were compiled*
Information was received from 15 States and 2 Canadian provinces
and was arbitrarily divided Into 4 geographical areas:
northwest, southern states, mid-Atlantic, and northeast. A total
of 3,695 collform values and 3,574 fecal collform values were
Included 1n the tabulations. The prime objective was to
determine the correlation between the two Indicator groups and
secondarily,to determine whether .or not collform data could be
used as a basis for evaluation of a potential fecal coliform standard.
The data show that a 70 collform MPN per 100 ml at the 50th
percentile was equivalent to a fecal•coliform MPN of 14 per 100 ml.
The data, therefore, indicate that a median value for a fecal
coliform standard is 14 ana tne 90th percentile shoulb not exceed 43
for a 5 tube, 3 dilution metnod (Hunt and Springer, 1973).
Evaluation of the microbiological suitability of waters
for recreational taking of shellfish should be based upon the
fecal coliform bacterial levels. When possible,
samples should be collected under those conditions of tide and
reasonable rainfall when pollution 1s. most likely to be maximum
in the area to be classified. The median fecal collform value
should not exceed an MPN of 14 per 100 ml and not more than 10
percent of the samples should exceed an MPN of 43.
-------
REFERENCES CITED
Berg, G. 1474. Reassessment of the virus problem 1n sewage and
In surface and renovated waters. Sixth International Water
Poll. Res. California. Pergamon Press.
Bonde, G.J., 1966. Bacteriological methods for the estimation of
water pollution. Hlth. Lab. Sc1. 3:124.
Bonde. G.J., 1974. Bacterial Indicators of sewage pollution.
International symposium on Discharge of Sewage from Sea
Outfalls. Pergamon Press.
Butler, P., e_t aj[., 1972. Test, monitoring and Indicator
organisms* In: A Guide to Marine Pollution, E.D. Goldberg, ed.
Gordon and Beach, N.Y.
Datta, N. and J. Olearte, 1974. R-Factors 1n strains of
Salmonella typh.1 and Shlgel la dysenterl J[ Isolated during
epidemics 1n Mexico: Classification by compatibility.
Antimicrobial Agents and*Chemotherapy 5: 310.
Gallagher, T.P., e_t a_l_., 1969. Pollution affecting shellfish
harvesting 1n Mobile Bay, Alabama, Tech. Servs. F.W.P.C.A,
Southeast Hater Lab., Athens, Georgia.
-------
Celclrelch, E.E., e_t aj_. , Ifrn. Fecal -col1forn-organ1 sm medium
for the membrane filter technique. Amer. Water Uorks Assn.,
51:208.
Geldrelch, E.E. and B.A. Kenner, 196Q. Concepts of fecal
streptococci 1n stream pclli'tion. ^our. Water Poll. Contr.
Fed. 41: R 33F.
Geldrelch, E.F., 1°72. Tuffalo Laboratory recreational water
quality: A study in bacterlol oqlcal data Interpretation.
•'.later Res. 6: PI 2.
Geldrelch, E.E., 1971. M1crohioloq1cal criteria concepts for
coastal bathing waters. ncoan Mnt. (In Press*).
Reldrelch, E.E., 1974. V.oritorini narlne waters for
microbiological quality. WPO Conference - Scientific Aspects
of Marine Pollution, r.enev^ Sv-/1 tzerland.
Guarraia, L. .1. 1972. Brief literature review of Klehslella as
«
pathogens. In s^^insr on t!ie glnni^icance of Pecal collforns
in Industrial Waste. E.P.T. T. P.. 3, 'lat^owal C1eld
Investinations Center, Oenvor, Colo. n. ^4.
Hunt, n.A. and J. Springer, 1974. Prel1n1nary report on "A
comparison of total collforn and fecal coll-forn values 1n
-------
.shellfish growing areas and a proposed fecal collform growing
area standard. Presented at 8th National Shellfish
Sanitation Workshop. (FDA. Wash. D.C.).
Janssen, W.A., 1974. Oysters; Retention and excretion of three
types of human water-borne disease bacteria. Health Lab.
Sc1. 11:20.
Kelly, C.B. and W. Arclsy, 1954. Survival of enteric organisms
1n shellfish. Pub. KeaUh Repts. 69:1205.
Kovacs, N., 1959. Enteric fever 1n connection with pollution of
seawater. Western Australia. Report of the Commissioner of
Public Health for the year 1958.
Metcalf, T.G. and W.C. Stiles, 1968. Viral pollution of
shellfish 1n estuary waters. Jour. San. Eng. Dlv. Proc. Amer.
Soc. Civ. Eng. 94:595.
McCoy, J.H., 1964. Salmonella. 1n crude sewage, sewage effluent
and sewage polluted natural waters. Abs. 1n Water Poll. Res.
pp 206.
Moore,'B. 1959. Sewage contamination of coastal bathing waters
In England and Wales. Jour. Hyg., 57:435.
97
-------
Moore, B., 1971. The Health hazards of pollution 1n mlcroblal
aspects of pollution. Sykes and Skinner, Eds., Academ. Press.
London: pp 11-32.
Olearte, 0. and E. Galindo, 1973. Salmonella typhl resistance to
chloramphenlcal, amph1c1111n and other antimicrobial agents:
Strains Isolated during an extensive typhoid fever epidemic
1n Mexico. Antimicrobial Agents and Chemotherapy 4:597.
Plotkln, S.A. and M. Katz, 1967. Minimal Infective doses 'of
viruses for man by the oral route, in.: Transmission of Viruses
by the Water Route. G. Berg, Ed., John Wiley Intersdence,
N.Y. p. 155.
Presnell, N, 1974. Discussion of fecal conforms for shellfish
growing waters. Proc. 7th_ National Shellfish Sanitation Workshop
Oct. 21 - 22, 1971. Ratcliff and Wilt, Eds., FDA. Wash., D.C.
Reltler, R. and R. Sellgaman, 19*57. Pseudomonas aerug-jnosa 1n
*
drinking water. Jour. Appl. Bact. 20:145.
Rlngem, L.M. and C.H. Drake, 1952. A study of the Incidence of
Pseudomonas aeruginosa fron various national sources. Jour
Bact. 64:841.
-------
Scarplno, D., 1974. Human enteric viruses and bacterlophages as
Indicators of sewage pollution. International Symposium from
Sea Outfalls. Pergamon Press.
Sears, H.J. e_t_ aj_., 1950. Persistence of Individual strains of
Escherchla coll in the Intestinal tract of man. Jour. Bact.
59:293.
Shuval, H.J., ejt aK, 1971. Natural 1nact1vat1on processes
of viruses in seawater. Proc. Natural Specialty
Conf. on Disinfection. Amer. Soc. of C1v1l Engineers, N.Y.
Slanetz, L.W., e_t aj_., 1965. Correlation of collform and fecal
streptococci Indices with the presence of Salmonella and
enteric viruses 1n sea water and shellfish. Adv. 1n Water
Pollution Res. 2nd International Confr., Tokyo. 3:17.
Smith, R.S., e_t aj_., 1951. Bathing water quality and health. I.
Great Lakes (U.S. Public HeaJI th Service, Cincinnati, Ohio).
Smith, R.S. and T.D. Woolsey, 1952. Bathing water quality and
health. II. Inland river and pool. (U.S.,Public Health
Service, Cincinnati, Ohio).
99
-------
Smith, R.S. and T.D. Woolsey, 1961. Bathing water quality and
Public Health. III. Coastal Waters (U. S. Public Health
Service, Cincinnati, Ohio).
Snow, D.J.R., 1959. Typhoid and City Beach. Western Australia.
Report of the Commissioner of Public Health for the Year
1958, p. 52.
Stevenson, A.H., 1953. Studies of bathing water quality and
health. Amer. Jour. Public Health 43:529,
*
Taylor, E.W., 1968. Forty-second .report of Director of Water
Examination, 1965-66, Metropolitan Wa'ter Board, London, p.
117.
Van Donsel, D.J. and E.E. Geldrelch, 1971. Relationships of
Salmonel lae to fecal conforms 1n bottom sediments. Water
Res. 5:1079.
Wolf, H.W., 1972. The collform count as a measure of water
quality. In Water Pollution Microbiology. R. Mitchell, Eds.,
WHey-Inter science.
-------
COLOR
CRITERIA;
Waters shall be virtually free from substances producing
Sffjectionable 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
ctmpounds originating from the decomposition of naturally-occurring organic
matter (AWWA, 1971). Sources of organic material include humic materials from
the soil such as tannins, humic acid and humatesj 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 humic materials (Sawyer,
1960). Natural color is reported in color "units" which generally are determined
by use of the platinum-cobalt method (Standard Methods, 1971).
tot
-------
There is no general agreement as to the chemical composition of natural
color, and in fact,the composition may vary chemically from place to place
(AWWA, 1971). Black and Christman (1963a) characterized color-causing colloids
examined as aromatic, polyhydroxy, methoxy carboxylic acids. Shapiro (1964)
characterized color-causing constituents as being dialyzable and composed
of aliphatic, polyhydroxy, 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 m>L 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 aesthetic1
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 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.
102.
-------
The effects of color on public water supplies also are principally
aesthetic. The 1962 Drinking Water Standards (PHS, 1962) reooranended that
color in finished waters should not exceed 15 units on the platinumr-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. Photo-
synthetic 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 shortwave or blue end of the spectrum (Birge and
Juday, 1930). In 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 photo-
synthetically 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 require-
ments (NTAC, 1968). The depth at which such a compensation point is reached,
called the 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 surtirearized '(NAS, 1974)
Table 2 lists the maximum value used as a source of water for various industries
and industrial uses. Through treatment, such waters can be made to meet almost
any industrial requirement.
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
-------
REFERENCES CITED;
American Water Works Association, 1971. Water quality and treatment.
3rd edition, McGraw-Hill Book Co., New York.
B1rge, E.A. and C. Juday, 1930. A second report on solar radiation and
Inland lakes. Trans. Wise. Acad. Science, Arts, Let., 25: 285.
Black, A.P. and R.F. Christman, 1963a. Characteristics of colored surface
waters. Jour. Amer. Water Works Assn., 55: 753.
Black, A.P. and R.F. Christman, 1963b. Chemical characteristics of fulvlc
adds. Jour. Amer. Water Works Assn., 55: 897.
National Academy of Sciences/ National Academy of Engineering, 1974.
Water quality criteria, 1*72. U.S. Government Printing Office,
Washington, D. C.
National Technical Advisory Committee to the Secretary of the Interior, 1968.
Water quality criteria. U.S. Government Printing Office, Washington, D. C.
Public Health Service, 1962. Drinking water standards, 1962. PHS Publ. No. 956,
U.S. Government Printing Office, Washington, D. C.
Sawyer, C.N., 1960. Chemistry for sanitary engineers* McGraw-Hill Book Co.,
New York.
.Shapiro, J., 1964. Effect of yellow organic adds on Iron and other metals 1n
water. Jour. Amer. Water Works Assn., 56: 1062.
-------
Standard Methods for the Examination of Water and Wastewater, 13th Edition, 1971
American Public Health Assn., Washington, D. C.
Welch, P.S., 1952. Limnology. McGraw-Hill Book Co., Inc., New York.
-------
COPPER
CRITERIA;
1.0 mg/1 for domestic water supplies (welfare).
For freshwater and marine aquatic life, 0.1 times
a 96-hour LC50 as determined through nonaerated
bioassay using a sensitive aquatic resident species.
JHTHpDUCTION:
'topper occurs as a natural or native metal and in various mineral
forms such as cuprite and malachite. The most important copper ores
civ sulfides, oxides, and carbonates. Copper has been ndned and used
in a variety of products by man since prehistoric times. Uvs for copper
include electrical products, coins, and metal platirg. Coppc" frequently
ir, alloyed with other metals to form various biasses and bronzes. Oxides
r.nd sulfates of copper are used for pesticides, algdcides, and fungicides.
Copper frequently is incorporated into paints ard wood preservatives to
inhibit growth of algae and invertebrate organisms, such as the woodborer,
Teredo_, on vessels.
Copper is an essential trace element for the propagation of plants
tnd performs vital functions in several enzymes and a major role in the
synthesis of chlorophyll. A shortage of copper in soil may lead to
chlorosis which is characterized by yellowing of plant leaves. In
copper deficient soils it may be added as a trace nutrient supplement to
other fertilizers.
107
-------
Copper is required in animal metabolism. It is important in
invertebrate blood chemistry and for the synthesis of hemoglobin. In
some invertebrate organisms a protein, hemocyanin, contains copper and
serves as the oxygen-carrying mechanism in the blood. An overdose of
ingested copper in mammals acts as an emetic.
In examining over 1500 surface vater samples fron the United States,
Kopp and Kroner (1967) found soluble copper in 71* percent of the samples
vf.th an average concentration of 15 Wg/1 and a maximum concentration of
?.80 Mg/1 of copper. The average concentration of copper in seawater
approximately 3.0 yg/1 (Mero, 196U).
RATIONALE.;
Concentrations of copper found in natural waters are not known to
have an adverse effect on humans. Prolonged oral administration of
excessive quantities of copper may result in liver damage, but water
supplies seldom have sufficient copper to effect such damages. Young
children require approximately 0.1 mg/day of copper for normal growth and
the daily requirement for adults was estimated to be about 2 mg/day
(Sollman, 1957). Copper in excess of 1 mg/1 may impart some taste to
water. Because of a possible undesirable taste in drinking water at
higher concentrations, a limit of 1 mg/1 is recommended.
precipitated copper in lake bottom muds resulting from
copper sulfate application to control nuisance algae, Mackenthun and
Cooley (1952) concluded that the toxic limit to a midge, Tendipi de s plumo sus ,
-------
and a fingernail dam, Pisidium idahoense was about 9,000 mg/kg of copper
in mud on a dry weight basis .
toxicity of copper to aquatic life is dependent on the alkalinity
of the water as the copper ion is complexed by anions present, which in turn
affect toxicity. At lower alkalinity copper is generally more tor.J.c to
aquatic life. Other factors affecting toxicity include pH and organic
compounds. Relatively high concentrations of copper may be tolerated by
adult fish for short periods of time; the critical effect of copper appears
to be its higher toxicity to young or juvenile fish.
Jtoudoroff and Katz (1953), in reviewing literature on the acute
toxicity of copper, concluded that in most natural fresh waters in the
United States copper concentrations below 25 yg/1 as copper evidently
are not rapidly fatal for most common fish species. In acute tests
Jones (1964 ) reported that copper sulfate in soft water (12 mg/1 CaCX^)
was toxic to rainbow trout at 60 yg/1 copper. In very hard water
(320 mg/1 Ca003) the toxic concentration was 600 yg/1 copper. A summary
of acute toxicity data is given in Table 3. In checking this table the
reader should consider the species tested, pH, alkalinity, and hardness
if alkalinity is not given (in most natural waters alkalinity parallels
hardness). In general the salmonids are very sensitive and the
centrarchids are less sensitive to copper.
-------
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References Cited in Table 3:
1. Benoit, D.A. (1975). Chronic effects of copper on survival, growth,
and reproduction of the bluegill Lepomis macrochirus. Crans. Amer. Fish.
Society, 104:353.
2. Brown, V.B., 1968. The calculation of the acute toxicity of mixtures
of poisons to rainbow trout. Water Res. 2:723.
3. Brungs, W.A., et al_., 1973. Acute and long-term accumulation of copper
by the brown bullhead Ictalurus nebulosus. Jour. Fish. Res. Bd. Can.,
30:583.
4. Brungs, W.A., e_t al_-, 1974. Acute and chronic toxicity of copper to the
fathead minnow (Pimephales promelas Rafinesque) in a surface water
of variable water quality. (In press).
5. Calamari, D., and R. Marchetti,'1973. The toxicity of mixtures of metals
and surfactants to rainbow trout. Water Res. 7:1453.
6. Chapman, 6.A. Unpublished data available at the National Water Quality
Laboratory, Duluth, Minnesota.
7. Doudoroff, P., 1956. Some experiments on the toxicity of complex
cyanides to fish. Sewage Ind. Wastes. 28:1020.
8. Holland, G.A., e_t al_., 1960. Toxic effects of organic pollutants on young
salmon and trout. State of Washington Dept. of Fisheries Res. Bull.
No. 5. 264 p.
119
-------
9. McKim, J.M., and D.A. Benoit, 1971. Effects of long-term exposures of
copper on survival, growth, and reproduction of brook trout
Salvelinus fontinalis. Jour. Fish. Res. Bd. Can. 28:655.
10. Mount, D.I., 1968. Chronic toxicity of copper to fathead minnows
(Pimephales promelas Rafinesque). Water Res. 2:215.
11. Mount, D.I., and C.E. Stephan, 1969. Chronic toxicity of copper to
the fathead minnow (Pimephales promelas) in soft water. Jour. Fish.
Res. Bd. Can.. 26:2449.
12. National Water Quality Laboratory (NWQL), 6201 Congdon Boulevard,
Duluth, Minnesota (Unpublished data).
13. Pickering, Q.H., and C. Henderson, 1966. The acute toxicity of some
heavy metals to different species of warm water fishes. Int. Jour. Air
and Water Poll.,-10':453.
14. Rehwoldt, R., et al_., 1971. Acute toxicity of copper, nickel, and zinc
ions to some Hudson River fish species. Bull. Environ. Contam.
Toxicol. 6:445.
15. Rehwoldt, R., et al_., 1972. The effect of increased temperature upon the
acute toxicity of some heavy metal ions. Bull. Environ. Contam.
Toxicol. 8:91.
16. Sprague, J.B., 1964. Lethal concentrations of copper and zinc for young
Atlantic salmon. Jour. Fish. Res. Bd. Can. 21:17.
-------
17. Tarzwell, C.M., and C. Henderson, 1960. Toxicity of less common metals
to fishes. Ind. Wastes. 5:12.
18. Trama, F.B., 1954. The acute toxicity of some common salts of sodium,
potassium, and calcium to the common bluegill (Lepomis macrochirus
Rafinesque). Proc. Acad. Nat. Sci., Philadelphia. 106:185.
19. Turnbull, H«» ejt a]_., 1954. Toxicity of various refinery materials
jo freshwater fish. Ind. Eng. Chem. 46:324.
20. Wallen, I.E., et afL, 1957. Toxicity to Gambusia af finis of certain pure
chemicals in turbid water. Sewage Ind. Wastes. 29:695.
21. Wellborn, Jr., T.L., 1969. The toxicity of nine therapeutic and
herbicidal compounds to striped bass. The Prog. Fish. Cult. 31:27.
22. Wilson, R.C.H., 1972. Prediction of copper toxicity in receiving
waters. Jour. Fish. Res. Bd. Can. 29:1500.
Ill
-------
Sprague and Ramsay (1965) determined the level beyond which the
organism can no longer survive for an indefinite period of time for
juvenile Atlantic salmon, Salmo salar, at 17° C. For copper in a water
with a total hardness of Ik mg/1 as CaC03 this level was 32 yg/1. At
28 yg/1 copper there were no salmon deaths in 168 hours.
Sprague (19610 found that Atlantic salmon, Salmo salar. tended to
avoid a concentration of copper as low as k.O yg/1.
Mount and Stephen (1969) reported that in chronic tests with fathead
minnows in water with a hardness of 200 mg/1 as CaCOa, 33 Wg/1 copper
did not affect survival or the physical appearance of the fish "but did
prevent spawning. No effects were noted at lU.5 Pg/1 copper. In soft
water with a hardness of 30 mg/1 as CaCO , the no-effect level for the
•3
fathead minnow, Pimephales prccnelas, was about 10.6 ug/1 copper. These
investigators reported application factors of between 0.13 and 0.22 and
between 0.03 and 0.08, respectively, for soft and hard water, 'the
maximum concentration of copper having no detectable effect on the brown
bullhead, Ictalurus nebulosus, in 600 days in water with a hardness of
202 mg/1 as CaC03 was determined to be between 16 and 27 yg/1 (Brungs et_ &!,..._,
1973). Benoit (1975) exposed bluegills (Lepomis macrochirus) to copper in
soft water for 22 months and found the "no-affect" level to be 21 ug/1
copper.
McKim and Benoit (1971) conducted chronic tests with brook trout,
Salvelinus fontinalis. exposed to copper in water with a mean alkalinity
of Ul.6 mg/1 as CaCO3. A copper concentration of 17.5 yg/1 did not
adversely affect survival, growth, or spawning of adult brook trout or
the hatchability of the eggs; however this concentration affected the
survival and growth of juveniles. The "no-effect" level established
122
-------
for the young brook trout was 9.5 yg/1 cbpper. In a second generation
exposure of brook trout to copper, McKim and Benoit (197*0 found that
exposure to "sublethal concentrations of copper from yearling through
spawning to 3 month juveniles" was sufficient to establish a "no-effect"
concentration (i.e., the "no-effect" level noted above caused no
adverse effects on the second generation). These authors reported
an application factor of between 0.17 and 0.10 of the 96-hour LC50 value
for the brown trout only.
Biesinger and Christensen (1972) found a 16 percent reproductive
impairment at 22 yg/1 copper for Daphnia magna in chronic (3-veek) tests
in water with a total hardness of U5.3 mg/1 as CaC03. Ihe 3-week LC50
was kk yg/1 copper. The total copper concentration having no effect on
Campeloma decisum, Physa integra, and Gammarus pseudolimnaeus in chronic
studies was between 8.0 and lU.8 yg/1 in water with a total hardness of
^5.3 mg/1 as CaC03 (Arthur and Leonard, 1970).
The concentration of copper that has been associated experimentally
with no harmful effect for several aquatic species is about 5 to 15 yg/1.
This is very close to the average ambient freshwater concentration now
found where copper occurs in measurable quantity. In waters with high
alkalinity and/or with much organic material many species will be able
to tolerate higher ambient copper concentrations. In such cases, the
criterion should not exceed 0.1 of the 96-hour LC50 (the approximate
mean application factor from tests reported above) as determined
through bioassays using sensitive resident species.
Copper is present in sea water at a concentration of approximately
3 yg/1 but copper added to the marine environment is readily precipitated
in the alkaline and saline environment. Toxicity of copper to fishes in
123
-------
aarine waters has not been studied, but for Nereis virens. a polychaete
invertebrate, the toxic threshold for copper vas 100 yg/1 (Raymont and
Shields, 196U). Copper is toxic to oysters at concentrations above
100 yg/1 (Galtsoff, 1932). dendenning and North (i960) found
inhibition of photosynthesis in the giant kelp, Macrocystis pyrifera,
at copper concentrations of 60 yg/1. This commercially important marine
plant is used for several industrial processes and for important food
additives. In areas where this plant is especially significant it may
be prudent to establish a restrictive copper criterion.
Mult softshell clans, Mya arenaria, were the most sensitive
marine macroorganisms tested in static copper toxicity bioassays.
ICO, LC50 and LC100 values after 168 hours at 30 o/oo salinity and
22°C were 25, 35 and 50 ug/1, respectively. At 17°C, these values
were 75, 86 and 100 ug/1, respectively, for the same time period
(unpublished manuscript). Copper is selectively concentrated over
zinc by adult softshell clams, Mya arenaria, Concentrations of
greater then 20 ug/1 are fatal after exposure of several weeks
(Pringle, et al., 1968). The 9-day LC50 for newly hatched Fundulus
heteroclitus larvae was 160 ug/1 (Gentile, 1975).
These data are insufficient to derive a satisfactory criteria
number, but it is apparent that copper does exhibit toxicity to the
few species tested. Therefore, it is recommended that .1 of the
96-hour LC50 for a sensitive aquatic species present be adopted.
-------
The only known industrial use of water affected by copper is the
production of textiles which requires a minimum concentration (as low as
10 pg/1) for some select processes. If needed, the low concentrations
found in natural water may be reduced readily by various forms of treatment
including coagulation and precipitation or ion exchange resins. Thus, no
water quality criterion for copper in industrial water supplies is proposed.
The minimum reported concentration of copper that begins to exhibit
toxicity to some agricultural plants is 100 yg/1, which is considerably
more than the average found in the Nation's waters. The adverse effect
of copper on plants can be overcome readily, by proper management through
irrigation or by the addition of materials such as lime, phosphate.
fertilizers, or iron salts to the soil (Reuther and Labanauskas, 1966).
No criterion is proposed for copper in water used for agricultural
purposes.
Oopper sulfate has been used widely in the control of algae in
water supply reservoirs and in recreational lakes. At Madison, Wisconsin,
its use for such purposes began in 1918 (Mackenthun and Oooley, 1952).
Copper sulfate was used first a.c a fish poison in 191^ and in 1953 it
was used experimentally by the Massachusetts Division of Fisheries and
Game in accelerating fish movements in ponds which were being fyke-trapped
for the removal of overabundant pan and weed species. It has since
been used as a standard accessory tool in netting operations (Tompkins and
Bridges, 1958).
-------
REFERENCES CITED:
Arthur, J.W. and E.N. Leonard, 1970. Effects of copper on Gammarus pseudolijnnaeus,
Physa Integra and Campeloma in soft water. Jour. Fish. Res. Bd. Canada,27:1277
Benoit, D.A., 1975. Chronic effects of copper on survival, growth, and reproduction
of the bluegill Lepomis macrochirus. Trans. Amer. Fish. Society, 104:353.
Bieslnger, K.E. and 6.M. Christensen, 1972. Effects of various'metals on survival,
growth, reproduction, and metabolism of Daphnia maqna. Jour. Fish. Res. Bd.
Can., 29: 1691.
Brungs, W.A., et al., 1973. Acute and long-term accumulation of copper by the brown
bullhead Jctalurus nebulosus. Jour. Fish. Res. Bd. Canada, 30:583.
Clendenning, K.A. and W.J. North, 1960. Effect of wastes on the giant kelp,
Macrocystis pyrifera. In: Proceedings, 1st International Conference on waste
Disposal in the Marine Environment. Pergamon Press, N.Y.
Cohen, J.M., e_t al_., 1960. Taste threshold concentrations of metals in
drinking water. Jour. Amer. Water Works Assn., 52: T299.
Doudoroff, R. and M. Katz, 1953. Critical review of literature on the toxicity
of industrial wastes and their components to fish, II: The metals, as salts.
Sew. and Ind. Wastes, 25: 802.
Eisler, R., Environmental Research laboratory, Narragansett, R.I., Toxicities of
salts of metals mixtures to marine bivalve molluscs. Unpublished manuscript.
Galtsoff, P.A., 1932. Life in the ocean from a biochemical point of view. Jour.
Washington Academy Sciences, 22:246.
12.5-
-------
Gentile, J., 1975. Environmental Research Laboratory, Narragansett, R.I., Semi-
Annual Report.
Jones, J.R.E., 1964. Fish and river pollution. Butterworths, London.
Kbpp, J.P. and R.C. Kroner, 1967. Trace metals in waters of the United States,
October 1, 1962 to September 30, 1967. U.S. Dept. of Interior, Fed. Water
Poll. Control Admin., Cincinnati, Ohio.
Mackenthun, K.M. and H.L. Codley, 1952. The biological effect of copper
sulphate treatment on lake ecology. Wise.Acad. Sci. Arts Lett., 41: 177.
McMrn, J.M. and D.A. Benolt, 1971. Effects of long-term exposures-to copper
on survival, growth and reproduction of brook trout (Salvelinus fontinails).
Jour. Fish. Res. Bd. Can., 28: 655.
Mero, J.L., 1964. Mineral resources of the sea. American Elsevler Publishing
Company.
Moore, G.T. and K.F. Kellerman, 1905. Copper as an algidde and disinfectant
in water supplies. U.S. Dept. of Agriculture, Bureau of Plant Industry,
Bull. No. 76, U.S. Govt. Print. Office, Washington, D. C.
Mount, D.I. and C.E. Stephan, 1969. Chronic toxiclty of copper to the fathead
minnow (Pimephales promelas) in soft water. Jour. Fish. Res. Bd. Can.,
26: 2449.
Pringle, B.H., et al., 1968. Trace metal accumulation by estuarine mollusks.
Jour. Sanit. Eng. Div., 94SA3:455.
-------
Raymont, J.E.G. and J. Shields, 1964. Toxlclty of copper and chromium In th«
marine environment, In: Pearson, E.A. (ed), advances In water pollution
research, proceedings, 1st international conference. Macmlllan Co., N.Y.
Reuther, W. and C.K. Labanauskas, 1966. Copper, In: Chapman, H.D., ed.,
diagnostic criteria for plants and soils. Univ. Calif., Berkley.
Soilman, T.H., 1957. A manual of pharmacology, 8th edition. W.B. Saunders
Company, Philadelphia, Pennsylvania.
Sprague, J.B., 1964. Avoidance of copper-zinc solutions by young salmon In tht
laboratory. Jou.r. Water Poll. Cont. Fed., 36: 990.
Sprague, J.B. and B.A. Ramsay, 1965. Lethal levels of mixed copper-zinc
solutions for juvenile salmon. Jour. F1sh. Res. Bd. Canada, 22: 425.
Tompkins, W.A. and C. Bridges, 1958. The use of copper sulfate to increase
fykenet catches. Progressive Fish-Culturist , 20:16
Warnick, S.L. and H.L. Bell, 1969. The acute toxicity of some heavy metals
to different species of aquatic insects. Jour. Fed. Water Poll. Cont.
Fed., 41:(Part 1), 280.
127
-------
CYANIDE
CRITERIA:
5.0 ug/1 for freshwater and marine aquatic
life and wildlife.
INTRODUCTION:
Cyanide is one of the simplest and most readily formed organic
moieties. Cyanide and compounds of cyanide are almost universally
present where life and industry are found. Besides being very important
in a number of manufacturing processes, they are found in many plants
and animals as metabolic intermediates whicn generally are not stored
for long periods of time.
In addition to the simple hydrocyanic acid (HCN), the alkali
metal salts such as potassium cyanide (KCN) and sodium cyanide (NaCN),
are commonly occurring forms and sources of cyanide. The latter
compounds are readily dissolved in water; the extent of HCN formation
is pH-dependent. A significant fraction of the cyanide exists as HCN
molecules up to a pH of approximately 8, and the fraction increases
rapidly as the pH of the solution decreases. When these simple salts
dissociate in aqueous solution, the cyanide ion combines with the
hydrogen ion to form hydrocyanic acid, which is toxic to aquatic life.
Chemically, the cyanide ion behaves similarly to the halide ions-
chloride, fluoride, bromide and iodide.
fZB
-------
The cyanide ion combines with numerous heavy metal ions to form
metallocyanide complexes. The stability of these anions is highly
variable. Those formed with zinc and cadmium are not stable; dissocia-
tion and production of hydrocyanic acid in near neutral or acidic
environments is rapid. In turn, some of the metallocyanide anions are
extremely stable. Cobaltocyanide is difficult to destroy with highly
destructive acid distillation in a laboratory. The iron cyanides are
also very stable but exhibit the phenomenon of photodecomposition, and
in the presence of sunlight the material dissociates to release the cyanide
ion, thus affecting toxicity; at night the reaction may reverse to produce
a less toxic form or state.
A wide variety of organic compounds may contain cyanide functional
groups. These compounds belong to a class of organic chemicals called
nitriles, few of which dissociate to liberate cyanide ions or molecular
HCN. In addition, there are also complex organic acids, alcohols, esters,
and amides that contain the cyanide radicals. These organic compounds
are used for numerous products or may be a waste by-product. Their
toxicity, persistence, and chemistry in the aquatic environment are not
well known except for a few specific compounds.
Cyanide toxicity is essentially an inhibition of oxygen metabolism,
i.e., rendering the tissues incapable of exchanging oxygen. The
cyanogen compounds are true noncumuiative protoplasmic poisons (can
be detoxified reaiily) since they arrest the activity of all forms of
animal life. Cyanide shows a very specific type of toxic action. It
-------
inhibits the cytonhrome oxidase system which facilitates electron transfer
from reduced metabolites to molecular oxygen. The ferric iron-porphyrin
molecule responsible for the catalytic action of cytochrome oxidase is
the reactive site where cyanide combines with ferric(+++) iron atoms to
form a reversible complex. Other enzymes containing a metal porphyrin
molecule, e.g., peroxidases and xanthine oxidase, are also strongly
inhibited by cyanide. Only undissociated HCN inhibits the consumption
of oxygen in the tissues, causing cellular asphyxia (histotoxic anoxia)
by attaching itself to the iron of the prosthetic group of the enzyme
cytochrome oxidase.
Hydrocyanic acid can be absorbed readily and carried in the
plasma but does not combine with hemoglobin because its iron atom is
divalent (ferrous). Instead, cyanide combines with methemoglobin, a
mildly oxidized form of hemoglobin in which the iron atom is trivalent
(ferric). Methemoglobin, which cannot carry oxygen, normally repre-
sents only a small fraction of the total hemoglobin. Since it forms an
irreversible and innocuous complex with cyanide, it is an active cyanide
detoxifying agent. Amyl nitrite and other agents can be used to
increase the level of methemoglobin to counteract cyanide toxicity.
A few of the ways in which cyanide can be metabolized within a pattern
of normal physiology are by the production of thiocyanate, reaction
with hydroxocobalamin to form the harmless cyanocobalamin, combination
with ami no acids, oxidation to carbon dioxide and formate, etc. The
conversion of only free cyanide and not organically bound cyano groups
(20
-------
to thiocyanate (SCN~) by action of the enzyme rhodanese 1s considered
to be the primary method of detoxification of cyanide. Rhodanese is
absent from blood and skeletal muscle, but 1s abundant in the liver.
Thiocyanate is eliminated irregularly and slowly in the urine.
The action of cyanide on the respiration of the cell and the
primary methods of detoxification of cyanide have been noted above.
However, it should be pointed out that cyanide does not completely
abolish cellular respiration. It is possible that a small amount of
residual respiratory activity is made possible by cytochrome b activity,
since this substance does not require the cyanide-susceptible cyto-
chrome oxidase. An alternative explanation of residual respiratory
activity of the cyanide-poisoned system is found in the action of the
flavin aerobic dehydrogenases, which can transfer hydrogen to molecular
oxygen without the cytochrome system.
The persistence of cyanide in water 1s highly variable. This
variability is dependent upon the chemical form of cyanide in the
water, the concentration of cyanide, and the nature of other
constituents. Cyanide may be destroyed by strong oxidizing agents
such as permanganates and chlorine. Chlorine is commonly used to
oxidize strong cyanide solutions to produce carbon dioxide and ammonia;
if the reaction is not carried through to completion, cyanogen chloride
may remain as a residual and this material is also toxic. If the pH
'of the receiving waterway is add and the stream is well aerated,
-------
gaseous hydrogen cyanide may evolve from the waterway to the
atmosphere. At low concentrations or toxicity and with acclimated
microflora, cyanide may be decomposed by microorganisms in both
anaerobic and aerobic environments or waste treatment systems.
Hydrocyanic acid probably is the most toxic form of cyanide in
water. The ratio of hydrocyanic acid-to total cyanide is quite
variable. Under natural conditions this variation is due to fluctuations
in pH. Photochemical action can also affect this ratio. Fluctuation
in pH is caused by acid wastewater discharges,and photosynthetic and
respiration cycles of aquatic plant life affect the formation, stability
and toxicity of HCN. Since such chemical and physical conditions will
dictate the form of cyanide, the cyanide criteria must be based on the
concentration of total cyanide present in the water.
RATIONALE:
Cyanide ingested by humans at quantities of 10 mg or less per day
is not toxic and is biotransformed to the less toxic thiocyanate.
Lethal toxic effects from the ingestion of water containing cyanide
occur only when cyanide concentrations are high and over-
whelm the detoxifying mechanisms of the human body. Continuous long-
term consumption of up to nearly 5 mg per day has shown no injurious
effects (Bodansky and Levy, 1923).
-------
A review of the available pertinent data on the acute toxicity
of simple cyanides to fish reveals the following-miniinum.lethal {threshold)
concentrations of free cyanide from data obtained from experiments
ranging from 12 minutes to 10 days: brook trout, Salvelinus
fontinalis (Karsten, 1934); rainbow trout, Salmo gairdneri (Herbert and
Merkens, 1952); brown trout, Sal mo trutta (Burdick, e_t al_., 1958);
small mouth bass, Micropterus dolomieu (Burdick, et^a]_., 1958);
bluegills, Lepomis macrochirus (Dc-udoroff, ejt al_., 1966); and fathead
minnows, Pimephales promelas (Doudoroff, 1956), are reported to be
50, 70 (60 determined concentration), 70, 104, 150, and 180 ug/1 as
cyanide, respectively.
Research at the University of Minnesota has revealed that the
minimum lethal (threshold) concentrations, as determined from continuous
flow bioassays in which routine analyses for cyanide were performed,
are generally lower than the above reported values. The threshold
concentrations, expressed as rng/1 cyanide, for the brook trout, Salvelinus
fontinalis; bluegill, Lepomis macrochirus; and fathead minnow, Pimephales
promelas, were determined to be 0.057 at 10°C, 0.104 at 25°C, and 0.120
at 25°C, respectively (Broderius, 1974).
In review it can be concluded that free cyanide concentrations
in the range from 50 to 100 ug/1 as cyanide have proven eventually
133
-------
fatal to many sensitive fishes,and levels much above 200 ug/ll
probably are rapidly Tatal to most fish species.
Downing (1954), Cairns and Scheier (1958), and Burdick, e_t al_.
(1958) have shown that the toxicity of free cyanides increases with any
reduction in dissolved oxygen below the 100 percent saturation level.
Cairns and Scheier (1958) observed that even periodic lowering of
dissolved oxygen decreased the tolerance of bluegills to cyanide.
The tolerance of fish to cyanide solutions that are rapidly lethal
has been reported to decrease with a rise in temperature. This increased
toxicity may be explained in part by the increased metabolism of fish
at higher temperatures.
Contradictory information from the literature indicates the
uncertainty between the relationship of toxicity of simple cyanides
to fish and the pH of the test solution. However, since undissociated
hydrogen cyanide has been demonstrated to be the toxic cyanide species
in simple cyanide solutions, changes in the pH of natural waters below
a value of about 8.3 should have no measurable effect on the acute
toxicity of simple cyanides to fish. There is no apparent relationship
among toxicity to fish, the alkalinity and the hardness of the dilution
water.
Cyanide is acutely toxic to most fishes at concentrations ranging
from 50 to 200 ug/1 (Herbert and Merkens, 1952; Burdick, ejtal_., 1958;
Cairns and Scheier, 1958; Doudoroff, 1956; Turnbull, e_t al_., 1954;
Lipschuetz and Cooper, 1955; Washburn, 1948).
-------
Some Information on chronic or sublethal effects of cyanide is also
available. Leduc (1966) found increased intestinal secretions in
the fish, Cichlasoma bimaculatum, at concentrations as low as 20 ug/1 and
reduced swimming capability at concentrations of 40 ug/1. Costa (1965)
reported that three common species of fish detected and avoided cyanide
concentrations of 26 ug/1 in approximately an hour or less. Exposure
to cyanide concentrations as low as 10 ug/1 reduced the swimming ability
or endurance of brook trout, Salvel inus f ont ina1is (Neil, 1957).
Growth, or food conversion efficiency of coho salmon, Oncorhynchus
kisutch, was reduced at hydrogen cyanide concentrations of 20 ug/1.
Small freshwater fish of the family Cichlidae exposed to a cyanide
concentration of 15 ug/1 lost weight more rapidly than the control fish
in water free from cyanide (Leduc, 1966).
The effects of cyanide on marine life have not been investigated
adequately to determine separate water quality criteria, but based on
the physiological mechanisms of cyanide, toxidty to marine life
probably is similar to that of freshwater life. Since marine waters
generally are alkaline, the toxicity of cyanide should be less than in
fresh waters where pH fluctuations occur more readily and frequently.
Thus, an additional factor exists to provide a margin of safety and
compensation for a lack of specific data on which to base the criterion
for marine aquatic life.
Cyanide has not been reported to have any direct effect on
recreational uses of water other than its effects on aquatic life. No
information is available on adverse effects of cyanide in agricultural
-------
practices nor in Industrial uses of water containing cyanide.
Since cyanide concentrations as low as 10 ug/1 have been reported
to cause adverse effects on fish, an ambient concentration of 5 ug/1
is believed to provide protection with a reasonable margin of safety.
-------
HEFERENCES CITED
Bodansky, M., and M.D. Levy, 1923. I: Some factors Influencing the
detox1cation of cyanides 1n health and disease. Arch. Int. Med. 31: 373.
Broderius, S.J., 1974, Testimony in the matter of proposed toxic pollutant
effluent standards for Aldr1n-Dleldr1n, et aJL Fed. Water Poll. Cont.
Agency (307) Docket No. 1.
Burdlck, G.E., _e_t_al_., 1958. Tox1c1ty of cyanide to brown trout and small mouth
bass. New York F1sh and Game Jour., 5: 133.
Cairns, J., Jr. and A. Scheler, 1958. The effect of periodic low oxygen
upon toxldty of various chemicals to aquatic organisms. Proc. 12th
Industr. Waste Conf., Purdue Univ. Eng. Bull., 42: 165.
Costa, H.H., 1965. Responses of freshwater animals to sodium cyanide
solutions. Ceylon Jour. Sic., Biol. Sic., 5: 41.
Doudoroff, P., 1956. Some experiments on the toxldty of complex cyanides
to fish. Sew. and Ind. Wastes, 28: 1020.
Doudoroff, P.».£t aj.., 1966. Acute toxldty to fish of solutions containing
complex metal cyanides, 1n relation to concentrations of molecular hydro-
cyanic add. Trans. Amer. Fisheries Soc., 95: 6.
Downing, K.M., 1954. The Influence of dissolved oxygen on the toxlclty of
potassium cyanide to rainbow trout. Jour. Exp. B1ol.t 31: 161.
Herbert, D.W.M. and J.C. Merkens, 1952. The toxldty of potassium cyanide
to trout. Jour. Exp. Biol., 29: 632.
137
-------
Karsten, A., 1934. Investigation of the effect of cyanide on Black
Hills trout. South Dakota School of Mines Black Hills Engineer, 22:145.
Leduc, H., 1966. Some physiological and biochemical responses of fish
to chronic poisoning by cyanide. Ph.D. Thesis, Oregon State University,
Corvallis, Oregon.
Lipschuetz, M. and A.L. Cooper, 1955. Comparative toxicities of potassium
cyanide and potassium cuprocyanide to the western black-nosed dace
(Rhinichthys atraUrius meleagris). New York Fish and Game Jour., 2:194.
Neil, J.H., 1957. Some effects of potassium cyanide on speckled trout,
Salvelinus fontinalus. 4th Ontario Industrial Wastes Conference,
Water Poll. Advisory Committee, Ontario Water Resources Commission, p. 74-96.
Turnbull, H., et al_., 1954. Toxicity of various refinery materials
to freshwater fish. Ind. and Eng. Chem., 46:24.
Washburn, G.H., 1948. The toxicity to warm water fishes of certain
cyanide plating and carburizing salts before and after treatment
by the alkali-chlorination method. Sew. Works Jour., 20:1074.
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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.
-------
Tne standard method of analyzing for gases in. solutions
has been the ,yan Slyke method (Van Slyke, e_t al., 1934) ; nov7 gas
chromatography also is used for determination of
individual and total gases. For determination of total gas
pressure, 'Veiss has developed the saturometer, a device
based upon a thin-wall silicons rubber tube th^t is
permeable to gases but impermeable to watery gases pass from
the water through the tube, thus raising the internal gas
pressure which is measured by a manometer or pressure gauge
connected to the tube (NA3, 1974). This method alone does
not separate the total gas pressure into the separate
components, but "//inkier 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 one hundred
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 presures is very low. Dissolved
nitrogen (l^) / wnich comprises roughly 30 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 (rJebeker, £t al., 1976a.).
Percent saturation of water containing a given amount of
gas varies with the absolute temperature and with the
pressure. Jecause of the pressure changes, percent
saturation with a given amount of gas changes with death of
the water. Gas sunersaturation decreases by 10 percent per
meter of increase in water depth due to hydrostatic
pressure; a gas that is at 130 percent saturation at the
Ml
-------
surface would be at 100 percent saturation at 3 meters1
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 due to 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 aJL., 1975). Cool, gas-saturated water is heated as it
passes through the condenser or heat exchanger. As the
temperature of the water rises, percent saturation increases
due to 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; and 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)i detection and avoidance of gas concentrations by
fish, intermittent exposure of fish to gas concentrations,
and bioassays of many species of fish exposed to different
concentrations of gas. Several conclusions resulting from
these studies are:
(1) When either juvenile or adult salmonids are
confined to shallow water (1 m), substantial mortality
occurs at and above 115 percent total dissolved gas
saturation.
(2) When either juvenile or adult salmonids are free to
sound and obtain hydrostatic compensation either in the
laboratory or in the field, substantial mortality still
occurs when saturation levels (of total dissolved gases) exceed
120 percent saturation.
-------
(3) On the basis of survival estimates made in the
Snake River from 1966 to 1975, it is concluded that juvenile
fish losses ranging from 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.
(U) 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 non-salmonids.
I-/*"
-------
Dawley and Ebql (1975) found that exposure of juvenile
spring chinook salmon, Qncorhynchus t shawyjzscha, and
steelhead trout, Salmo gairdneri, to 120 percent saturation
for 1.5 days resulted in over 50 percent mortalityj 100
percent mortality occurred in less than 3 days. They also
determined that the threshold level where significant
mortalities begin occurring is at 115 percent nitrogen
saturation (111 percent total gas saturation in this test).
Rucker (1974) , using juvenile coho salmon, Oncorhynchus
ki sutch, determined the effect of individual ratios of
oxygen and nitrogen and established that a decrease in
lethal effect occurred when the nitrogen content fell below
109 percent saturation even though total gas saturation
remained at 119 percent saturation, indicating the
importance of determining the concentration of the
individual components (0 and N_) of the atmospheric
supersaturation. Nebeker, et al. (1976a), using juvenile
sockeye salmon, Oncorhynchus nerka, also showed that there
was a significant increase in fish mortality when the
nitrogen concentration was increased while holding the total
percent saturation constant. They also showed that there
-------
was no significant difference in fish mortality at different
CO„ concentrations.
Research completed by Bouck, e_t al. (1975) showed that
gas supersaturated water at and above 115 percent total gas
saturation is acutely lethal to most soecies of salmonidg,
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, Qncoriiynchus nerka. Chinook salmon, Oncorhy nchus_
tshawyts^cna, were intermediate in sensitivity. Coho salmon,
Oncorhyncaus kisutch, were significantly the more tolerant
of the salmonids though still much more susceptible than
non-salmonids like bass or carp.
Daphnia magna exhibited a sensitivity to supersaturation
similar to that of the salmonids (Nebeker, et ajL., 1975),
with 115 percent saturation lethal within a few days;
stoneflies exhibited an intermediate sensitivity similar to
bass with mortality at 130 percent saturation; and crayfish
-------
10
were very tolerant, with levels near 140 percent total gag
saturation resulting in mortality.
-------
11
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
eyew 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°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.
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12
It has been shown by Bouck, et al. (1975) and Dawley, et
al. (1975) that survival of salmon, and steelhead smalts 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. 1976b).
-------
13
.REFERENCES CITED;
Bouck, G.R., et al., 1975. Mortality, saltwater adaptation
and reproduction of fish exposed to gas supersaturated
water. Unpublished report. U.S. Environmental
Protection Agency, Western Fish Toxicology Station,
Corvallis, Oregon.
Clay, A., et al., 1975. Experimental induction of gas bubble
disease in menhaden. Presented at the American Fisheries
Society, September, 1975, Las Vegas, Nevada. New England
Aquarium, Boston, Mass.
Dawley, E.M., and W.J. Ebel. 1975. Effects of various
concentrations of dissolved atmospheric gas on
juvenile Chinook salmon, Oncorhynchus tshav
-------
14
Dawley, E., et al., 1975. Bioassays of total dissolved
gai pressure. Unpublished report. National Marine
Fisheries Service, Seattle, Washington.
De*tont, J.D. and R.W. filler. 1972. First reported
incidence of gas bubble disease in the heated
effluent of a steam electric generating station.
Proc. 25th Annual Meeting, Southeast Assoc. of Game
and Fish Commissioners.
Ebel, W.J., et al., 1975. Effect of atmospheric gas
supersaturation caused by dams on salmon and
steelhead trout of the Snake and Columbia Rivers.
Final Report. Northwest Fisheries Center, NMFS,
Seattle, Washington.
Keup, L.E. 1975. Factors in fish kill investigations.
tfater and Sewage Works, 121:48.
-------
15
Lindroth, A, 1957. Abioqenic qas supersaturation of
river water. Arch. Hydrobiology, 53:589.
Malouf, R., 1972. Occurrence of gas bubble disease
in three species of bivalve mollusks. Jour. Fish.
Res. Bd. Can., 29:588.
Marcello, R.A., et al., 1975. Evaluation of alternative solutions to
gas bubble disease mortality of menhaden at Pilgrim Nuclear
Power Station. Yankee Atomic Electric Co., Westboro,
Mass. YAEC-1087.
National Academy of Sciences, National Academy pf
Engineering. 197^. Water quality criteria, 1972.
U.S. Government Printing Office, Washington, D.C.
Nebeker, A.V., et al., 1975. Effects of gas supersaturated water
on freshwater invertebrates. Proc. Gas Bubble Disease
Workshop. Battelle Northwest, ERDA Special Report (In Press).
-------
Nebeker, A.V., et al., 1976a. Nitrogen, oxygen, and carbon
dioxide as factors affecting fish survival in gas supersaturated
water. Trans, toner. Fish. See. (In Press).
Nebeker, A.V., et al., 1976b. Survival of coho and sockeye
salmon molts in seawater after exposure to gas supersaturated
water. Trans. Amer. Fish. Soc. (In PresS).
Renfro, W.C. 1963. Gas bubble mortality of fishes in
Galveston 3ay, Texas. Trans. Amar. Fish. Soc.
32:320.
Rucker, R.R. 1974. Gas bubble disease: Mortalities of
coho salmon, Oncorhynchus kisutch, in water with
constant total gas pressure and different
oxygen-nitrogen ratios. National Oceanic cind
Atmo-s. \dmin., -^atl. Mar. Fish. Serv., Northwest
Fish Center, Seattle, Washington, unpublished manuscript.
ML 8
-------
17
Rulifson, R.L. and G. Abel. 1971. Nitrogen
supersaturation in the Columbia and Snake Rivers.
Tech Kept. TS-09-70-208-016,2, Environmental
Protection Agency, Region X, Seattle, Washington.
Van Slyke, D.D., et al« 1934, Studies of gas and
electrolyte equilibria in blood,.XVIII. Solubility
and physical state of atmospheric nitrogen in blood
cells and plasma. Jour. Biol. Chenu , 105:571.
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HARDNESS
INTRODUCTION;
Water hardness is caused by the polyvalent metallic ions dissolved in
water. In fresh waters these are principally 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 carbonate (CaCO ).
The concept of hardness comes from water supply practice. It is
measured by soap requirements for adequate lather formation and as an
indicatonof 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).
Classification of Water by Hardness Content
Cone., mg/1 CaC03 Description
0-75 soft
75 - 150 moderately hard
150 - 300 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
Y47
-------
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. Since bicarbonates generally are measured as
alkalinity, the carbonate hardness usually 1S 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 chemical 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 (NTAC, 1968) and NAS (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 o* the
principal cations contributing to hardness, e.g., calcium, or a combination of both
-------
effects. Stiff (1971) presented a theory (without proof) that if cupric ions
were the toxic form of copper while 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 LC5Q 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/1 as CaC03 (NAS, 1974).
Limits on hardness for industrial uses are quite variable. Table 4 lists
maximum values that have been accepted by various industries as a source of raw
water (NAS, 1974). Subsequent treatment generally can reduce hardness to tolerable
limits although costs of such treatment are an important factor in determining
its desirability 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
than hardness.
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Table 4
Maximum Hardness Levels Accepted
by industry as a Raw Water Source*
Maximum Concentration
Industry mg/1 as CaC(h
Electric Utilities 5,000
textile 120
Pulp and Paper 475
Chemical 1,000
Petroleum 900
Primary Metals 1,000
* Requirements for final use within a process may be essentially
zero which requires treatment for concentration reductions.
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REFERENCES CITED'.
Doudoroff, P. an4 Katz, M., 1953. Critical review of literature on the
toxicity of industrial wastes and their components to fish. II. The
metals as salts. Sew. and Ind. Wastes, 25, 7, p. 802.
National Academy of Sciences,, National Academy of Engineering, 1974. Water
.quality criteria, 1972. U.S. Government Printina Office, Washinaton, D. C,
National Technical Advisory Committee to the Secretary of the Interior, 1968.
Water quality criteria. U.S. Government Printing Office, Washington, D. C.
Sawyer, C.H., 1960. Chemistry for sanitary engineers. McGraw Hill Book
Co., Inc., New York.
Stiff, M.J., 1971. Copper/bicarbonate equilibria in solutions of bicarbonate
ion at concentrations similar to those found 1n natural water. Water
Research, 5: 171.
is/
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IRON
CRITERIA:
0.3 mg/1 for domestic water supplies (welfare).
l.O^mg/1 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; especially this is true 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+'f), and the ferric, or trivalent (Fe*4"4;)
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 pumoed or drained. For
practical purposes the ferric (Fe444") form is insoluble. Iron can exist in
natural organometallic or humic compounds and colloidal forms. Black or
brown swamp waters may contain iron concentrations of several mg/1 in the presence
or absence of dissolved oxygen, but this iron form has litte effect on aquatic
/5K
-------
life because it is complexed or relatively inactive chemically or
physiologically.
In stratified lakes with anaerobic hypolimnia, soluble ferrous iron
occurs in the deep, anaerobic waters. During the autumnal or vernal overturns
and with aeration of these lakes, it is oxidized rapidly to the ferric ion
that precipitates to the bottom sediments as a hydroxide, Fe(OH)3, or with
other anions. If hydrogen sulfide (H2S) is present in anaerobic bottom
waters or muds, ferrous sulfide (FeS) may be formed. Ferrous sulfide is a
black compound and results in the production of black mineral muds.
Prime iron pollution sources are industrial wastes, mine drainage
waters, and iron-bearing groundwaters. In the presence of dissolved oxygen,
iron-in water from mine drainage is precipitated as a hydroxide, Fe(OH)3.
These yellowish or ochre precipitates produce "yellow boy" deposits found
in many streams draining coal mining regions of Appalachia. Occasionally
ferric oxide ^6203) is precipitated forming red waters. Both of these
precipitates form as gels or floes that may be detrimental, when suspended
in water, to fishes and other aquatic life. They can settle to form flocculant
materials that cover stream bottoms thereby destroying bottom-dwelling
invertebrates, plants or incubating fish eggs. With time these floes can
consolidate to form cement-like materials, thus consolidating bottom gravels
into pavement-like areas that are unsuitable as spawning sites for nest
building fishes; particularly this is detrimental to trout and salmon
populations whose eggs are protected in the interstices of gravel and
incubated with oxygen-bearing waters passing through the gravel.
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RATIONALF.:
Iron is an objectionable constituent in water supplies for either domestic
or industrial use. Iron appreciably affects the taste of beverages (Riddlck,
ejt aj_., 1958) and can stain laundered clothes and plumbing fixtures. A study
by the Public Health Service (Cohen, et al_., 1960) indicates that the taste
of iron may be detected readily, at 1.8 mg/1 1n spring water and at 3,4 mg/1
1n distilled water.
The dally nutritional requirement for Iron 1s 1 to 2 mg, but Intake of
larger quantities is required as a result of poor absorption. Diets contain
7 to 35 mg per day and average 16 mg (Sollman, 1957). The Iron criterion 1n
water 1s fso prevent objectionable tastes or laundry staining (0.3-mg/1)
constitutes only a small fraction of the Iron normally consumed and 1s of
aesthetic rather than toxicologlcal significance.
Warnick and Bell (1969) obtained 96-hour LC$Q values of 0.32mg/1 1r0rt
for mayflies, stoneflies, and caddisflies; all are important fish food organisms*
Brandt (1948) found iron toxic to carp, Cyprlnus carplo, at concentrations of
0,9 mg/1 when the pH of the water was 5.5. Pike, Esox lucius. and trout
(species not known) died at iron concentrations of 1 to 2 mg/1 (Doudoroff
and Katz, 1953). In an iron polluted Colorado stream, neither trout nor other fish-
were found until the waters were diluted or the iron had precipitated to
effect a concentration of less than 1.0 mg/1 even though other water quality
constituents measured were suitable for the presence of trout (FWPCA, 1967).
Ferric hydroxide floes have been observed to coat the gills of white perch
Roccus americanusl minnows and silversides, Menidia sp. (Olsen, et al_.,
-------
1941). The smothering effects of settled Iron precipitates may be particularly
detrimental to fish eggs and bottom-dwelling fish "food organisms. Iron deposits
1n the Brule River, Michigan and Wisconsin wen; found to have a residual long-
term adverse effect on fish food organisms even after tne pumping of iron-
bearing waters from deep shaft iron mines had ceased (West, et^ aJL, 1963).
Settling Iron floes have also been^reported to trap and carry diatoms downward
in waters (01 sen, et a_l_., 1941).
Ellis (1937) found that in 69 of 75 study sites with good fish fauna,
the iron concentration was less than 10.0 mg/1. The European Inland .Fisheries
Advisory Commission (1964) recommended that iron concentrations not exceed
1.0 mg/1 in waters to be managed for aquatic life.
Based on field observations principally, a criterion of 1 mg/1 Iron
for freshwater aquatic life 1s believed to be adequately protective.
As noted, data obtained under laboratory conditions suggest a greater
toxicity for iron than that obtained in natural ecosystems. Ambient
natural waters will vary with respect to alkalinity, pH, hardness,
temperature and the presence of ligands which change the valence state
and solubility, and therefore the toxicity of the metal.
The effects of iron on marine life have not been investigated adequately
to determine a water quality criterion. Dissolved iron readily precipitates
1n alkaline sea waters. Fears have been expressed that these settled 1r6rt
floes may have adverse effects on Important benthic commercial mussels and
other shellfish resources.
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Iron has not been reported to have a direct effect on the recreational
uses of water other than Its, effects on aquatic life. Suspended troft precipitates
may Interfere with swimming and Be aesthetically objectionable. Deposits of'
ochre or reddish iron oxides can be aesthetically objectionable.
Iron at exceedingly high concentrations has been reported to be toxic to
livestock and Interfere with the metabolism of phosphorus (NA§, 1974),
Dietary supplements of phosphorus can be used to overcome this metabolic
deficiency (McKee and Wolf, 1963). In aerated soils, Iron 1n Irrigation Waters,
iis not toxic. Precipitated iron may complex phosphorus and molybdenum making
them less available as plant nutrients. In alkaline soils, Iron may be so
insoluble as to be deficient as a trace element and result 1n chlorosis* an
objectionable plant nutrient deffdency disease. Rhoades (1971) reported a
reduction 1n the quality of tobacco because of precipitated Iron oxides on
the leaves when the crop was spray.Irrigated with water containing S tog/1
of soluble iron.
For some Industries, iron concentrations in process waters lower than
that prescribed above for public water supplies are required or desirable.
Examples Include high pressure boiler feed waters; scouring, bleaching,
and dyeing of textiles; certain types of paper production; some chemicals;
some food processing; and leather finishing industries.
Iflo
-------
LITERATURE CITED:
Brandt, H.H., 1948. Intensified injurious effects on fish, especially
the increased toxic effect produced by a combination of sewage poisons.
Beitr. Mass. Abwass., Fischereichemi. 15.
Cohen, J. M., e£ aJL, 1960. Taste ^threshold concentrations of metals
in drinking water. Jour. Amer. Water Works Assn., 52: 660.
Doudoroff,- P., and M. Katz, 1953. Critical review of literature on the
toxicity of industrial wastes and their components to fish. II. The
metals, as salts. Sew. Ind. Wastes, 25:802.
Ellis, M. M., 1937. Detection and measurement of .stream pollution.
Bulletin, U. S. Bur. Fisheries, 48: 365.
European inland Fisheries Advisory Commissjon, 1964. Water quality criteria
for European freshwater fish. Report on finely divided solids and Inland
fisheries. Tech. Paper T.
FWPCA, 1967. Effects of pollution on the aquatic life resources of the
South Platte River Basin. Two volumes. South Platte River Basin
Project, Denver, Colorado and Technical Advisory and Investigations
Branch, Cincinnati, Ohio. Fed. Water Pollution. Cont. Admin., U.S.
Dept. of Interior.
McKee, J. £., and H. W. Wolf, 1963. Water quality criteria. State Water
Quality Control Board, Sacramento, California, Pub. 3-A.
157
-------
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972. U.S. Government Printing Office,
Washington, D. C.
Olson, R.A., et aj_., 1941. Studies of the effects of Industrial pollution
in the lower Patapsco River area. I: Curtis Bay Region,, Chesapeake
Biological Laboratory, Solomans Island, Maryland,
Rhoades, P.M., 1971. Relations between Fe in Irrigation water and lea*
quality of cigar wrapper tobacco. Agron. Jour., 63: 939.
Rlddick, T.M., et. al_., 1958. Iron and manganese in water supplies.
Jour. Amer. Water Works Assn., 50: 688.
Soilman, T.M., 1957. A manual of pharmacology, 8th ed. W.B. Saunders
Co., Philadelphia, Pa.
Warnlck, S.L. and H.L. Bell, 1969. The acute toxicity of some heavy
metals to different species of aquatic insects. Jour. Water Poll.
Cont. Fed., 41(Part 2): 280.
West, A.W., e_t a]_., 1963. Report on pollution of the interstate waters
. of the Menomlnee and Brule rivers, Michigan-Wisconsin. U.S. Dept.
of Health, Education and Welfare, Public Health Service, R.A. Taft
Sanitary Engineering Center, Cincinnati, Ohio.
-------
LEAD
CRITERIA;
50 ug/1 for domestic water supply (health);
0. 01 times the 96-hour LCc0 value, using the receiving
or comparable water as" the diluent and soluble lead
measurements (non-filtrable lead using an .0,45.
micron filter), for sensitive freshwater resident species.
INTRODUCTION;
In addition to their natural occurrence, lead and its compounds
may enter and contaminate the global environment at any stage during
mining, smelting, processing, and use. The annual increase in lead
consumption in the U. S. during the 10-year period from 1962-1971
averaged 2. 9 percent, largely due to increased demands for electro-
chemical batteries and gasoline additives (Ryan, 1971). In 1971 the
total U. S. lead consumption was 1, 431, 514 short tons, of which 42
percent came from recycled lead (Ryan, 1971). Of the 1971 U. S. lead
consumption, approximately 25 percent was as metallic lead or lead
alloy (Ryan, 1971; NAS, 1972). Non-industrial sources that may
contribute to the possibility of ingestion of lead by man include the
indoor use of lead-bearing paints and plaster, improperly glazed
earthenware, lead fumes on ashes produced in burning lead battery
casings, and exhaust from internal combustion engines.
Most lead salts are of low solubility. Lead exists in nature mainly
as lead sulfide (galena); other common natural forms are lead carbonate
(cerussite), lead sulfate (anglesite), and lead chlorophosphate
(pyromorphite). Stable complexes result also from the interaction of
ts9
-------
lead with the sulfhydryl, carboxyl, and amine coordination sites
characteristically found in living matter. The toxicity of lead in water,
like that of other heavy metals, is affected by pH, hardness, organic
materials and the presence of other metals. The aqueous solubility
of lead ranges from 500 ug/1 in soft water to 3 ug/1 in hard water.
Lead enters the aquatic environment through precipitation, lead
dust fallout, erosion and leaching of soil, municipal and industrial
waste discharges, and the runoff of fallout deposits from streets and
other surfaces. Extrapolations from recent studies (EPA, 1972;
University of Illinois, 1972) Indicate that nationally as imich as 5,000
tons,of lead per year may be added to the aquatic environment as a
result of urban runoff.
Mediterranean and Pacific surface waters contain up to 0. 20 and
0. 35 mg/ 1 of lead, respectively (NAS, 1972), which is about 10 times
the estimated pre-industrial lead content of marine waters. The lead
content of rivers and lakes also has increased in recent years (NAS,
1972). It may be inferred from available data that the mean natural
le'ad content of the world's lakes and rivers ranges from 1 to 10 ug/1
(Livingstone, 1963); the lead content of rural U. S. soils is 10 to 15
ug/g (Chow and Patterson, 1962), and the usual range of lead-in-soil
concentrations is 2 to 200 ppm, exclusive of areas near lead ore
deposits (Motto, ertaL, 1970), although many urban soil concentrations
are much higher.
In the analyses of over 1, 500 stream samples, Kopp and Kroner
report that lead was observed at measurable levels with a
Ito
-------
frequency of under 20 percent. The mean concentration of the positive
occurrences was 23 ug/1. The highest incidence of occurrence of lead
was observed in the Western Great Lakes Basin where the frequency
was slightly above 40 percent. The highest recorded concentration
was 140 ug/1 in the Ohio River at Evansville, Indiana.
RATIONALE;
As far as is known, lead has no beneficial or desirable nutritional
effects. Lead is a toxic metal that tends to accumulate in the tissues
of man and other animals. Although seldom seen in the adult population,
irreversible damage to the brain is a frequent result of lead intoxi-
cation in children. Such lead intoxication most commonly results from ingestion of
lead-containing paint still found in older homes. The major toxic
effects of lead include anemia, neurological dysfunction, and renal
impairment. The most common symptoms of lead poisoning are anemia,
severe intestinal cramps, paralysis of nerves (particularly of the arms
and legs), loss of appetite, and fatigue; the symptoms usually develop
slowly. High levels of exposure produce severe neurologic damage,
often manifested by encephalopathy and convulsions; such cases frequently
are fatal. Lead is strongly suspected of producing subtle effects (i. e.,
effects due to low level or long term exposures insufficient to produce
overt symptoms) such as impaired neurologic and motor development and
renal damage in children (EPA, 1973). Subclinical lead effects are
distinct from those of residual damage following lead intoxication.
Biochemical effects of lead include inhibition of erythrocyte delta-
aminolevulinic acid dehydrase (ALAD) activity, increased urinary
-------
excretion of delta-aminolevulinic acid (ALA-U), and increased blood
lead concentration;. The increase in ALA-U is of particular interest
because it rarely is produced by any substance other than lead; it is
related directly to inhibition of ALAD, the enzyme that converts ALA-U
to porphobilinogen. Recent work indicates that lead interferes with
heme biosynthesis, and thus elevates ALAHJ excretion, at levels well
below 40 ug/100 ml whole blood (Secchi, et aJU, 1974; Haeger-ftronsen,
et al., 1974). As a result of this work, the Center for Disease Control
has recommended that the upper limit of normal for lead in the blood
of children be revised downward from 40 ug to 30 ug/100 ml whole blood. An
ad hoc committee, appointed by the U. S. Public Health Service to
establish a daily permissible intake of lead without excessive body lead
burden in children, concluded that the level of such intake is 300 ug
from all sources (King, 1971). The gastrointestinal absorption and
retention of lead is greater in children than in adults, 53 percent and
18 percent, respectively, as shown in recent studies (Alexander, et
aJL , 1973). The average daily intake of lead from diet and air among
young children probably amounts to up to 2/3 of the daily permissible
intake, leaving a very narrow margin of safety (King, 1971; Lin-Fu,
1973). Compared to adults, food and air intake by children is
proportionally greater than their weight, e.g., a 1-year-old child,
with only about 1/7 of the body weight of an adult, has 1/4 to 1/3
of the daily adult air intake and 40 to 60 percent of the dietary intake
of an adult (NAS, 1972), so that his lead intake is proportionally
greater on a body weight basis. Alexander, et^aL (1973) have
suggested a daily limit of lead intake for 0-5 year-olds of 10 ug/kg.
-------
Considerable evidence has been developed demonstrating that laboratory
animals on high lead dosages show teratogenic effects; however, no such
effects have been observed in cattle or sheep (NAS, 1972). Epidemic -
logic studies have demonstrated no relationship between lead exposure
and cancer incidence in man, but it is known that lead at concentrations
of one percent or more in the diet causes renal cancer in rats (NAS, 1972).
The lead content in public water supplies in the U. S. in 1962
ranged from traces to 62 ug/1 (Dufor and Becker, 1964). Continuous
monitoring of the U. S. water supplies since 1962 has demonstrated
that their lead content has, in general, not exceeded the U. S. Public
Health Service standard of 50 ug/1 (USPHS, 1962). McCabe (1970)
reported on 2, 595 distribution samples and showed that 73 percent
contained less than the USPHS limit. McCabe, et^al. (1970) found
that the range of lead concentrations in finished U. S. community
water was from non-detectable to 0. 64 mg/1. Of the 969 water supplies
surveyed, 1. 4 percent exceeded 0. 05 nag/I of lead. Five of the water
supplies in this sample contained sufficient lead to equal or exceed
Kehoe's (1966) estimated maximum safe level of lead intake (600 ug/day)
without considering the possible additional contributions to the total
intake by other sources and routes of exposure. In drinking water lead
should be kept to a minimum; a criterion of 50 ug/1 is attainable and
protective. Experience indicates that fewer than four percent of the
water samples analyzed exceed the 50 ug/1 limit and that the majority
of these are due to corrosion problems and are not due to naturally
occurring lead content in raw waters.
763
-------
For the fathead minnow, Pimephales promelas, Pickering and
Henderson (196») have determined the 96-hour TLm values for lead
(chloride) to be 5. 6 to 7. 3 mg/1 in soft water (20 to 45 mg/1 as CaCO3>
Brown (1968) reported a 96-hour LCgQ of 1 mg/1 for rainbow trout,
Salmo gairdneri, in soft water (50 mg/1 as CaCO-j). The 96-hour
LCgQ value for lead in hard water was 482 mg/1 for fathead minnows
(Pickering and Henderson, 196#). Other short-term fish toxicity data
are in Table 5.
Preliminary information on 2- to 3-month exposures of rainbow and
brook trout indicated detrimental effects at 0.10 mg/1 of lead in soft
water (20-45 mg/1 as CaCC>3 ) (NAS, 1974). Growth of guppy species
was affected by 1. 24 mg/1 of lead (Crandall and Goodnight, 1962); Jones
(1939) and Hawksley (1967) found chronic or sublethal effects on three-
spine stickelback species from lead concentrations of 0.1 arid 0. 3 mg/1,
and the conditioned behavior of goldfish, Carassius auratus, in a light-
dark shuttlebox was adversely affected by 0. 07 mg/1 of lead in soft
water (50 ppm CaCC>3 added to deionized tap water) (Weir and Hine, 1970).
A concentration of 30 ug/1 lead in a 3-week exposure produced 16
percent reproductive impairment in Daphnia magna in water with a hard-
ness of 45 mg/1 CaCC>3 (Biesinger and Christensen, 1972). The 96-hour
LCgQ for rainbow trout species in hard water (alkalinity 243 mg/1) for
total lead was 471 mg/1 and the highest mean continuous flow concen-
tration that did not have an adverse effect on survival, growth, and
reproduction was 0.12 and 0. 36 mg/1 (Davies and Everhart, 1973).
-------
For dissolved lead, the 96-hour LCgn was 1. 38 mg/1 and the
no-effect level was 18 to 32 ug/1. Total and free lead were considered
to be the same in soft water. The 18-day LCgQ in soft water (alkalinity
26.4 mg/1) was 140 ug/1 and the highest mean continuous flow con-
centration that did not have an adverse effect on survival, growth
and reproduction was 6.0 to 11.9 ug/1. The no-effect concentrations
were determined on the occurrence of abnormal black tails caused
by chronic lead exposure. Acute toxicity values for several species
of fish in water of various qualities are presented in Table 5. When
referring to this table, the reader should consider the species tested,
pH, alkalinity, and hardness df alkalinity is not given (in most natural
waters alkalinity parallels hardness). In general, the salmonids are
most sensitive to lead in soft water, but the influence of pH, and
other factors on the solubility and form of the lead preclude the
recomendation of a freshwater criteria based on acute toxicities alone.
-------
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References Cited in Table 5:
1. Benoit, D. and G. Holcombe. Unpublished data available from
Water Quality Laboratory, Duluth, Minnesota.
2. Chapman, G. Unpublished data available from National Water
Quality Laboratory, Duluth, Minnesota.
3. Davies, P. H. and W. H. Everhart, 1973. Effects of chemical
variations in aquatic environments, Vol. III. Lead toxicity
to rainbow trout and testing application factor concept.
Ecological Research Series, EPA R3-73-011c.
4. Hawksley, R. A. 1967. Advanced water pollution analysis
by a water laboratory. Analyzer, 8:13.
5. Henderson, C. Unpublished data available from National
Water Quality Laboratory, Duluth, Minnesota.
6. Pickering, Q. H. and C. Henderson. 1965. The acute toxicity
of some heavy metals to different species of warm water fishes.
Purdue University. Proc. 19th Ind. Waste Conf., 578-591.
7. Tarzwell, C. M. and C. Henderson. 1960. Toxicity of less
common metals to fishes. Industrial Wastes, 5:12.
8. Turnbull, H., et_al. 1954. Toxicity of various refinery materials
to freshwater fish. Symposium of waste disposal in the
petroleum industry. Ind. Eng. Chem. 46:325.
9. Wallen, I. E., et^al. 1957. Toxicity to Gambusia affinis of
certain pure chemicals in turbid waters. Sewage and Industrial
Wastes, 29:695.
-------
There are few data from which to draw conclusions about the
relationship of safe levels for a given species in hard and soft water.
The differences in lethal levels between very soft and very hard water
is 100 to 500 times because of insolubility and precipitation. The
single source of information for chronic and acute effects in hard
and soft water is for rainbow trout (Davies and Everhart, 1973)
and shows these significant factors:
Hardwater data were reported in terns of both toted
•
and free lead. The static acute hard water bioassay
(mean hardness, 353.5 mg/1; methyl orange alkalinityr
243.1 mg/1; and pH, 8.02) demonstrated 96-hour TL
(50 percent tolerance, limit) xxmcentration of 471 mg/1
total lead and 1.38 mg/1 free lead.
In the chronic hard water tests, lead-attributed
mortalities occurred at levels of 3240 ug/1 total lead
and 64 ug/1 free lead, while the maximum acceptable
toxicant concentrations, based on the occurrence of
black tails, were found to be 120 ug/1 to 360 ug/1
total lead and 18 ug/1 to 32 ug/1 free lead.
Soft water data were reported in terms of free lead.
The flow-through acute soft water bioassay (mean hardness,
27.2 mg/1; methyl orange alkalinity, 26.4 mg/1; and
pH 6.88) demonstrated an 18-day TL5Q (50 percent tolerance
limit) of 140 ug/1 free lead.
171
-------
In the chronic soft water tests, leadiattributed
mortalities occurred in fry and fingerling fish at a
concsrttration of 95.2 ug/1, while three-year-old brood
.f&bh produced viable eggs and fry when exposed to any
but the highest (27.9 ug/1) concentration. The maximum
acceptable toxicant concentration, based on the occurrence
of black tails, was determined to be between 11.9 and 6.0 ug/1.
A criterion involving an application factor of 0. 01 multiplied by the
acute 96-hour LGjQ value expressed as dissolved lead is used to
estimate the safe concentration for various fish species. Based upon
the existing data, as well as upon the sensitivity of various species of
fish to other metals, it is highly probable that salmon, trout,
minnows and catfish will be especially sensitive to lead as compared
to bass, sunfish and perch. Therefore, tests for acute toxicity should
be performed on the more sensitive species when establishing standards
for lead. This approach requires the experimental determination of
LC CQ values before a criterion can be determined, but the extreme
effect of various water characteristics on lead solubility and toxicity
warrants this additional effort.
Berry (1924) found that a concentration of lead nitrate of 25 mg/1
was required to cause toxic effects to oats and tomato plants. At a
concentration of 50 mg/1, plant death occurred. Hopper (1937) found
that 30 mg/1 of lead in nutrient solutions was toxic to bean plants.
172-
-------
Wilkins (1957) found that lead at 10 mg/1 as lead nitrate reduced root
growth. Since dissolved lead contents in soils were usually from 0.05
to 5.0 mg/kg (Brewer, 1966), little toxicity can be expected. It was
shown that the principal entry of lead into plants was from aerial de-
posits rather than from absorption from soils (Page, et^aL , 1971),
indicating that lead that falls into the soil is not available to plants.
There is no question that some marine organisms can concentrate
the lead present in sea water. Wilder (1952) reported lobster dying in
6 to 20 days when held in lead-lined tanks. Calabrese, et al. (1973)
found a 48-hour LCsg of 1730 ug/1 and a 48-hour LCso of 245° US/1 for
oyster, Crassostrea virginica, eggs. The remarkable ability of the
eastern oyster, Crassostrea virginica, to concentrate lead was demon-
strated (Pringle, et^ al., 1968) by exposing them to flowing sea water
containing lead concentrations of 25 ug/1, 50 ug/1, 100 ug/1, and 200 ug/1;
after 49 days, the total accumulation of lead amounted to 17, 35, 75, and
200 ppm (wet weight), respectively, and those oysters exposed to the
two highest lead levels, upon gross examination, showed considerable
atrophy and diffusion of the gonadal tissue, edema, and less distinction
of hepatopancreas and mantle edge.
North and Clendenning (1958) reported that lead nitrate at 4.1 mg/1
of lead showed no deleterious effect on the photosynthesis rate in kelp,
Macrocystis pyrifera. exposed for four days. However/ there are
insufficient data, upon which to -base a marine criterion at this time.
-------
REFERENCES CITED
Alexander, F.W., et al_., 1973. The uptake and excretion by children of lead
and other contaminants, In: Proc. of the int'l. symposium on environmental
health aspects of lead. Amsterdam, October 2-6, 1972, Commission of the
European Communities, Luxembourg, p. 319.
Berry, R.A., 1924. The manurial properties of lead nitrate. Jour. Agric.
Sci. (London), 14: 58.
Biesinger, K.E., and G.M. Christensen, 1972. Effects of various metals on
survival, growth, reproduction and metabolism of Daphnia magna. Jour.
Fisb. Res. Bd. of Canada, 29: 1691.
Brewer, R.F., 1966. Lead, In: H.D. Chapman, Ed., Diagnostic criteria for
plants and soils. Berkeley, University of California, Division of
Agric. Sci., p. 213.
Brown, V.M., 1968. The calculation of the acute toxicity of mixtures of
poisons to rainbow trout. Water Res., 2: 723.
Calabrese, A., e£ al_., 1973. The toxicity of heavy metals to embryos of the
american oyster, Crassostrea virginica. Marine Biol., 18: 162.
Chow, T.J., and C.C. Patterson, 1962. The occurrences and significance of
lead isotopes in pelagic sediments. Geochim. Cosmochim. Acta., 26: 263.
Crandall, C.A. and C.J. Goodnight, 1962. Effects of sublethal concentrations
of several toxicants on growth of the common guppy, Lebistes reticulatus.
Limnol. Oceanog., 1: 233.
-------
Davles, P.H. and W.H. Everhart, 1973. Effects of chemical variations in aquatic
environments: volume III. Lead toxicity to rainbow trout and testing
application factor concept. Environmental Protection Agency, Ecol. Res. Series
Report, EPA-R3-73-011.
Durfor, C. and E. Becker, 1964. Selected data on public water supplies of the
100 largest cities in the United States, 1962. Jour. Amer. Water Works
Assn., 56: 237.
Environmental Protection Agency, 1972. Water pollution aspects of street
surface contaminants. EPA-R2-72-081, U.S. Environmental Protection
Agency, Washington, D.1 C.
Environmental Protection Agency, 1973. EPA's position on the health
implications of airborne lead. U.S. Environmental Protection Agency,
Washington, D. C.
Haeger-Aronsen, at a^., 1974. Effect of lead on delta-atninolevulinic acid
dehydratase activity in red blood cells. Arch. Environ. Health, 29:150.
Hawksley, R.A., 1967. Advanced water pollution analysis by a water
laboratory. Analyzer, 8: 13.
Hopper, M.C., 1937. Effect of lead on plants. Ann. Appl. Biol., 24: 690.
Jones, J.R.E., 1939. The relation between the electrolytic solution
pressures of the metals and their toxicity to the stickleback (Gasterosteus
aculeatus L.). Jour. Exp. Biol., 16: 425.
Kehoe, R.A., 1966. Under what circumstances is ingestion of lead dangerous?
In: Symposium on environmental lead contamination. U.S. Dept. of
Health, Education, and Welfare, Public Health Service, Publication 1440.
Washington, D.C., p. 51.
-------
King, B.G., 1971. Maximum daily intake of lead without excessive body lead-
burden in children. Amer. Jour. Dis. Children, 122:337.
Kopp, J.F., and R.C. Kroner, 1967. Trace metals in waters of the United
States. U.S. Department of the Interior, Federal Water Pollution Control
Administration, Cincinnati, Ohio.
Lin-Fu, J.S., 1973. Vulnerability of children to lead exposure and toxicity.
New Eng. Jour. Med., 289:1229.
Livingstone, D.A., 1963. Chemical composition of rivers and lakes, pp. G1-G64.
Geological Survey Professional Paper 440-G. In: M. Fleischer, Ed., Data of
Geochemistry (6th ed.). U.S. Government Printing Office, Washington, D.C.
McCabe, L.J., e_t al_., 1970. Survey of community water supply systems.
Jour. Amer. Water Works Assn. 62:670.
McCabe, L.J., 1970. Metal levels found in distribution samples. Amer.
Water Works Assn. Seminar on Corrosion by Soft Water, Washington, D.C. 9 pp.
Motto, H.L., et a!., 1970. Lead in soils and plants: Its relationship to
traffic volume and proximity to highways. Environmental Sci. Tech., 4:231.
National Academy of Sciences - Committee on Biologic Effects of Atmospheric
Pollutants, 1972. Lead: Airborne lead in perspective. The National Academy
of Sciences, Washington, D.C.
National Academy of Sciences, National Academy of Engineering, 1974. Water quality
criteria, 1972. U.S. Government Printing Office, Washington, D. C.
North, W.J. and K.A. Clendenning, 1958. The effects of waste discharges on
kelp. Annual Progress Report, Institute of Marine Resources (University of
California, LaJolla, California). IMR Reference 58-11.
-------
Page, A.L., et al_., 1971. Lead quantities in plants, soil and air near some
major highways in Southern California. Hilgardia, 41.
Pickering, Q.H., and C. Henderson, 1966. The acute toxicity of some heavy
metals to different species of warm water fishes. Int. Jour. Air-Water Poll.,
10:453.
Pringle, B.H., et al_., 1968. Trace metal accumulation by estuarine mollusks.
Jour. Sanit. Eng. Div.; Proc. Am. Soc. Civil Eng., 94:455.
Ryan, J.P., 1971. Minerals Yearbook, 1971. U.S. Department of the Interior,
Washington, D.C.
Secchi, G.C., et aL., 1974. Delta-aminolevulinic acid dehydratase activity
of erythrocytes and liver tissue in man. Arch. Environ Health, 28:130.
University of Illinois, 1972. Environmental pollution by lead and other metals
(NSF RANN Grant 61-31605), Progress Report, May 1 - October 31, 1972, Chapter 6.
University of Illinois at Urbana-Champaign.
USPHS, 1962. Public Health Service Drinking Water Standards, U.S. Department
of Health, Education, and Welfare, Public Health Service Publication 956
(Revised), Washington, D.C.
Weir, P.A., and C.H. Hine, 1970. Effects of various metals on behavior of
conditioned goldfish. Arch. Envir. Health., 20:45.
Wilder, D.C., 1952. The relative toxicity of certain metals to lobsters.
Jour. Fish Res. Bd. Can., 8:486.
Wilkins, D.A., 1957. A technique for the measurement of lead tolerance in
plants. Nature, 180:37.
777
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MANGANESE
CRITERIA:
50 ug/1 for domestic water supplies (welfare);
100 ug/1 for protection of consumers of marine mollusks.
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
prgperly. 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.
176.
-------
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 ba.tteries (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, e_t al_., 1958; Illig, 1960). Consumer complaints arise
when manganese exceeds a concentration of 150 ug/1 in water supplies
(Griffin, 1960). These complaints are concerned'primarily with the brownish
staining of laundry and objectionable tastes in beverages. It is possible
that the presence of low concentrations of iron may intensify the adverse
effects of manganese. Manganese at concentrations of about 10 to 20 ug/1
is acceptable to most consumers. A criterion for domestic water supplies
of 50 ug/1 should minimize the objectionable qualities.
McKee and Wolf (1963) summarized data on toxicity of manganese to
freshwater aquatic life. Ions of manganese are found rarely at concen-
trations above 1 mg/1. The tolerance values reported range from 1.5 mg/1
to over 1000 mg/1. 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 concentrations of 2.2 to 4.1 mg/1, but permanganates are not
persistent because they rapidly oxidize organic materials and are
179
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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/1
(Fairbridge, 1966). The material is rapidly assimilated and bioconcen-
trated into nodules that are deposited on the sea floor. The major
problem with manganese may be concentration in the edible portions of
mollusks.as bioaccumulation factors as high as 12,000 have been reported
(NAS, 1974). In order to protect against a possible health hazard to
humans by manganese accumulation in shellfish, a criterion of 100 ug/1
is recommended for marine water.
Manganese is not known to be a problem in water consumed by livestock.
At concentrations of slightly less than 1 mg/1 to a few milligrams per
liter, manganese may 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/1 of manganese (NAS, 1974). But as stated above, manganese rarely
is found in surface waters at concentrations greater than 1 mg/1. 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/1 is suggested for consideration.
Most industrial users of water can operate successfully whare 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.
-------
REFERENCES CITED;
Fall-bridge, R.W. (ed), 1966. The encyclopedia of oceanography. Reinhold,
New York, New York.
Griffin, A.E., 1960. Significance and removal of manganese in water supplies.
Jour. Amer. Water Works Assn., 52: 1326.
Illig, G.L. Jr., 1960. Use of sodium hexametaphosphate in manganese stabilization.
Jour. Amer. Water Works Assn., 52: 867.
McKee, J.E. and H.W. Wolf, 1963. Water quality criteria. State Water Quality
Control Board, Publ. , Sacramento, California, Pub. 3-A.
National Academy of Sciences, National Academy of Engineering, 1974. Water
quality criteria, 1972. U.S. Government Printing Office, Washington, D. C.
»
Riddick, J.M., et al_., 1958. Iron and manganese in water supplies. Jour.
Amer. Water Works Assn., 50: 688.
Sollman, T.H., 1957. A manual of pharmacology. W.B. Saunders, Philadelphia,
Pennsylvania.
181
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MERCURY
CRITERIA:
2.0 ug/1 for domestic water supply (health);
0.05 ug/1 for freshwater aquatic life and wildlife;
0.10 ug/1 for marine aquatic life.
INTRODUCTION:
Mercury is a silver-white, liquid metal solidifying at -38.9° C
to form a tin-white, ductile, malleable mass. It boils at 356.9° C, has
a specific gravity of 13.6 and a vapor pressure of 1.2 x 10"3 mm of
mercury. Mercury has three oxidation states: (1) zero (elemental mercury);
(2) +1 (mercurous compounds); and (3) +2 (mercuric compounds). Mercury is
widely distributed in the environment and biologically is a non-essential
or non-beneficial element. Historically it was recognized to possess a
high toxic potential and was used as a germicidal or fungicidal agent for
medical and agricultural purposes.
Human poisoning by mercury or its compounds clinically has been
recognized. Although its toxic properties are well known, dramatic instances
of toxicosis in man and animals have occurred recently, e.g., the Minamata
Bay poisonings (Irukayama, et al., 1962; Irukayama, 1967). In addition
to the incidents in Japan, poisonings have also occurred in Iraq,
Pakistan and Guatemala as a result of ingestion of flour and seed treated
with methyl and ethylmercury compounds (Bakir, et al., 1973). Mercury Intoxi-
cation may be acute or chronic and toxic effects vary with the form of mercury
and its mode of entry into the organism. The mercurous salts are less
soluble than the mercuric and consequently are less toxic. For man, the
fatal oral dose of mercuric salts ranges from 20 mg to 3.0 g (Stokinger, 1963).
Symptoms of acute, inorganic mercury poisoning include pharyngitis, gastro-
enteritis, vomiting followed by ulcerative hemorrhagic colitis, nephritis,
183
-------
hepatitis, and circulatory collapse. Chronic mercury poisoning results from
exposure to small amounts of mercury over extended time periods. Chronic
poisoning from inorganic mercurials most often has been associated with
industrial exposure, whereas poisoning from the organic derivatives has been
the result of accidents or environmental contamination. Alkyl compounds are
the derivatives of mercury most toxic to man, producing illness, irreversible
neurological damage, or death from the ingestion of amounts in milligrams
(Berglund and Berlin, 1969).
The mercury content of unpolluted U.S. rivers from 31 States where
natural mercury deposits are unknown is less than 0.1 ug/1 (Wershaw,-1970).
Jenne (1972) found also that the majority of U.S. waters contained less
than 0.1 ug/1 of mercury. The lower limit of detection in these studies
was 0.1 ug/1. Total mercury values of 0.045 ug/1 recently were determined
in Connecticut River water by Fitzgerald and Lyons (1973) using more
sensitive methods. Marine waters have been shown to contain concentrations
of mercury from a low of .03 to a high of 0.2 ug/1, but most marine waters
fall within the range of .05 to .19 ug/1 mercury (Robertson, et aj_., 1972).
Mining, agriculture and waste discharges contribute to the natural levels
found.
RATIONALE:
Several forms of mercury, ranging from elemental to dissolved inorganic
and organic species, are expected to occur in the environment. The recent
discovery that certain microorganisms have the ability to convert inorganic and
organic forms of mercury to the highly toxic methyl or dimethyl mercury
has made any form of mercury potentially hazardous to the environment
/av
-------
(Jensen and Jernelov, 1969). In studies on the biochemical kinetics of
mercury methylation, Bisogni and Lawrence 0973) demonstrated that in
water,under naturally occurring conditions of pH and temperature, inorganic
mercury can be converted readily to methylmercury.
Wood (1974) has argued further that whenever mercury in any form is
added to the aquatic environment, a combination of microbially catalyzed
reactions and chemical equilibrium systems is capable of leading to steady
state concentrations of dimethyl mercury, methylmercuric ion, metallic
mercury, mercuric ion, and mercurous ion. Thus, it is evident that the
total mercury level should be the basis for a mercury criterion instead
of any particular form in which it may be found within a sample.
Hannerz (1968), using 0.1 mg/1 of several mercury compounds in ponds,
concluded that algae and other aquatic plants accumulate mercury primarily
by surface adsorption. This study demonstrated that all of the mercury
compounds used were taken up by fish both directly from the water and from
food. The accumulation rate was shown to be fast, while the elimination
rate was slow, leading to concentration factors of 3,000-fold and higher.
According to McKim (1974), concentration factors by fish in excess of
10,000 times the amount of mercury in the surrounding water have been
demonstrated.
In a test period of 20 to 48 weeks, several species of fish accu-
mulated more than 0.5 ug/gm mercury in their tissues from a water habitat
containing 0.018 to 0.030 ug/1 methy1mercury (MsKim, et al., in Press)
representing concentration factors of from 27,800 to 16,600.
-------
Quantities of ingested mercury safe for man can be estimated from
data presented in "Methyl Mercury in Fish" (1971). From epidemiological
evidence, the lowest whole-blood concentration of methy!mercury associated
with toxic symptoms is 0.2 ug/g. This blood concentration can be compared
to 60 ug/g mercury in hair. These values, in turn, correspond to prolonged
continuous exposure at approximately 0.3 mg/70 kg body weight/day. By
using a safety factor of 10, the maximum dietary intake from all sources,
including air, water, and food, should not exceed 30 ug/person/day mercury.
Although the safety factor is computed for adults, limiting ingestion by
children to 30 ug/day of mercury is believed to afford some lesser degree
of safety. If the exposure to mercury were from fish alone, the limit
would allow for a maximum daily consumption of 60 grams (420 g/week) of
fish containing 0.5 mg/kg mercury. A drinking water criterion of 2.0 ug/1
would permit a daily intake of 4 ug mercury assuming an average consumption
of 2 liters of water per day. If the mercury is not all in the alky! form,
a greater margin of safety will exist.
The levels of mercury in tissues of livestock consumed by humans should
not exceed 0.5 mgAg mercury. The tissue concentration of 0.5 mjAg correlates
approximately with a blood level of 0.1 ug/1. A mercury level of less than
1.0 ug/1 in livestock water is considered compatible with these concentrations,
thus a mercury concentration of 0,05-.ug/1 -will provide an adequate safety factor.
Several chronic toxicity tests have been conducted to measure the
adverse effects of organomercurials on survival, growth, and reproduction
of several fish species. In a 3-year chronic toxicity study involving
-------
three generations of brook trout, Salvelinus fontinails, exposed to
methylmercuric chloride, gross toxic symptoms were observed after six-
months' exposure of yearling trout to 2.9 ug/1 of mercury (McKim, et al.,
In Press). Spawning occurred at all lower mercury concentrations-
tested, but the offspring of parental fish exposed to 0.93 ug/1 of
mercury exhibited reduction in growth 90 days after hatching. After 24-
months' exposure of second generation fish to 0.93 ug/1, behavioral symptoms
were noted, there was no spawning, and mortality was 94 percent. No
adverse effects were observed in the brook trout exposed to methylmercuric
chloride concentrations of 0.29 ug/1 mercury and below. A full life cycle
chronic toxicity test was conducted with fathead minnows, Pimephales
promelas, exposed to methylmercuric chloride (Mount, 1974). All died after
three months' exposure to concentrations of 0.80 and 0.41 ug/1 mercury.
Ninety-two percent of the fish exposed to 0.23 ug/1 mercury also died within
the three-month period. Spawning was completely inhibited at 0.12 ug/1
mercury with males not developing sexually. No toxic effects were noted
on survival or growth of the offspring produced in 0.07 ug/1 of mercury.
Two chronic toxicity tests were conducted with one invertebrate,
Daphm'a magna. Mercury as mercuric chloride and methylmercuric chloride
caused significant reproductive impairment at concentrations of 2.7 ug/1
and 0.04 ug/1 mercury, respectively (Biesinger, 1974).
Matida, ejt al_. (1971) found that the LC$Q for phenylmercuric acetate,
methylmercuric chloride, and mercuric chloride with rainbow trout finger-
lings, Salmo gairdneri, were 8.5, 30, and 310 ug/1, respectively. Wobeser
(1973) examined the toxicity of methylmercuric chloride to two life stages
-------
of rainbow trout. The 96-hour 1X50 for newly hatched sac fry was 24 ug/1
of mercury, while rainbow trout fingerlings had a 96-hour LC50 value of 42 ug/1
Eisler (1974) found that concentrations of 1.0 ug/1 mercury represent
a distinct threat to selected species of marine organisms and, based on a
comparison with freshwater species, the accumulation of mercury is similar
in fresh and marine water.
Because of methylation and bi concentration of methyl mercury, mercury
limits must take into consideration the food chain transport path from
aquatic organisms to man. Regardless of the mercury form present, the
major portion of the mercury will ultimately reside in the bottom sediments
wherethrough microbial action, mono- and dimethyl mercury can be formed.
These forms of mercury are bioconcentrated many-fold in fish arid other
aquatic organisms because of the very rapid uptake and the relative inability
of the fish to excrete methylmercury from their tissues. As a result,
methyl mercury in fish tissues may exceed the 0.5 mgAg PDA guideline. This
occurs in water concentrations that have no observed toxic effects on the
fish. Methylation rates are highly dependent upon water quality conditions,
but sufficient evidence exists to suggest that the process can occur in the
pH range of 5 to 9 under aerobic or anaerobic conditions; hence, it is assumed
that methylation can and will occur in natural waters.
Dernethylation processes can deolete methylmercury concentrations in water.
Methylmercury aooears to nersist for sufficient time neriods, however to allow
>
uptake by aquatic organisms. Hence, demethylation orocesses can have an effect
on uptake rates of methylmercury, but do not terminate the transport path.
It appears that the methylation process takes place at the water/sediment
interface, particularly in the sediment area in which the benthic organisms
are most active. The movement of benthos within the sediments contributes
-------
to the methylation process by physically expanding the area of water/
sediment interface. Through ingestion of the detritus in the sediments,
benthos acquire a body burden of mercury that will in turn be trans-
ported to fish upon ingestion.
The FDA established guideline for mercury in edible fish is 0.5 vtg/kg.
Thus, the maximum level of mercury in receiving waters should be based on
the premise that this level should not be exceeded. A mercury concen-
tration factor for certain freshwater species has been shown to be in
excess of 10,000.
Upon dividing the 0.5 mg/1 FDA temporary tolerance by the
10,000-fold accumulation factor, a level of 0.05 ug/1 for total mercury in fresh water
results, which can be assumed to protect the human consumer of freshwater fish. It was
pointed out above that 0.04 ug/1 mercury as methylmercuric chloride caused
significant reproductive impairment to Daphnia magna. Recognizing, however,
that natural total mercury concentrations in fresh water are in the same
range, and that a total mercury level of 0.05 ug/1 would be divided among
several chemical forms which differ markedly in their toxicity, it is
believed that a criterion of 0.05 ug/1 total mercury will offer a reasonable
level of protection to freshwater aquatic life as well as to the human
consumer.
Recognizing that seawater contains about 0.1 ug/1 mercury and that
this level is 1/10 of that found by Eisler (1974) to represent a threat to
selected species of marine organisms, it is recommended that the criterion
for the protection of marine aquatic life be 0.1 ug/1.
119
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REFERENCES CITED
Bakir, P., S.F. Damluji, L. Amin-Zaki, M. Murtadha, A. Khalidi,
N.Y. Al Rawl, S. Tikriti, H.I. Dhahir, T.W. Clarkson, J.C. Smith
and R.A. Ddherty, 1973. Methylmercury Poisoning in Iraq. Science
181;230-240.
Berglund, F. and M. Berlin, 1969. Risk of accumulation In men and mammals and
the relation between bojiy burden of methyl mercury and toxic effects. In:
Chemical Fallout, Thomas, Springfield, 111. pp. 258-273.
Blesinger, K.E., 1974. Testimony in the matter of proposed toxic pollutant
effluent standards for Aldrin-Dieldrin, e_t al_. FWPCA (307) Docket No. 1,
Exhibit No. 14.
Bisogni, J.J. and A.W. Lawrence, 1973. Methylation of mercury in aerobic and
anaerobic environments. Technical Report 63, Cornell University Resources
and Marine Sciences Center, Ithaca, New York.
Elsler, R., 1974. Testimony in the matter of proposed toxic pollutant effluent
standards for Aldrin-Dieldrin, et al_. FWPCA (307) Docket No. 1.
Fitzgerald, W.F., and W.B. Lyons, 1973. Organic mercury compounds in coastal
waters. Nature, 242.
Food and Drug Administration, 1974. Poisonous or deleterious substances
in peanuts, evaporated milk, fish and shellfish. Proposed Rules,
Federal' Register. Dec. 6, 1974, Washington, D.C.
Hannerz, L., 1968. Experimental investigations on accumulation of mercury in
water organisms. Fishery Board of Sweden, Institute of Freshwater Research,
Drottinghoflw, Report 48.
-------
Irukayama, K., ejt a]_., 1962. Studies on the origin of the causative
agent of Minamata disease, III. Industrial wastes containing mercury
compounds from Minamata factory. Kumamoto Medical Jour., 15(2}:57.
Irukayama, K., 1967. The pollution of Minamata Bay and Minamata
disease. Adv. Water Poll. Res., 3:153.
Jenne, E.A., 1972. Mercury In waters of the United States, 1970-71. Open file
report, U.S. Geol. Surv., Menlo Park, California.
Jensen, S. and A. Jernelov, 1969. Biological methlyation of mercury 1n aquatic
organisms. Nature, 223.
Matida, Y., et_al_., 1971. Toxidty of mercury compounds to aquatic organisms
and accumulation of the compounds by the'organisms. Bull. Freshwater Res.
Lab., Vol. 21, No. 2.
McKim, J.M., 1974. Testimony in the matter 6f proposed toxic pollutant effluent
standards for Aldrln-Dieldrin, et al_. FWPCA (307) Docket No. 1.
McKim, J.M., «t ad., 1976. long term effect* of n*thyJ«excuric chloride
on three generations of brook trout (Salvelinus fontinalis);)toxicity,
accumulation, distribution, and elimination. Jour. Fish. Res. Bd.
Canada, (In Frees).
Methyl Mercury/1n fisH, 197/1. $ toxlcolog^c-e^d&nlologlc evaluation of
risks. Report from an expert group. Nord. Hyg. Tldskr., Suppl. 4.
-------
Mount, D.I., 1974. Chronic toxlclty of methylmercuric bh!6rfde*to fathead
minnow, left^mon^ln the matter of proposed toxic effluent standards for
Aldr1n-D1eldr1n, et'aK FWPCA (307) Docket No. 1, Exhibit No. 4.
Robertson, E.E., et al_., 1972. Battelle Northwest contribution-to the IDOE
Base-line Study, Battelle Northwest 1972 IDOE Workshop, p. 231.
Stoklnger, H.E., 1963. Mercury Hg237. In: Industrial Hygiene and Toxicology
Vol. 2, 2nd 4d. F.A. Patty, Ed., New York Inter science, p. 1090.
Hershaw, R.L., 1970. Mercury 1n the environment. Geological Survey Professional
faper #713., G.P.O.
Mobescr, 6.A., If73. Aquatic mercury pollution: Studies of Us occurrence and
pathologic effects on fish and mink. Thesis: Dept. of Vet. Pathology,
U. of Saskatchewan, Saskatoon, Canada.
Wood, J.M., 1974. Biological cycles for toxlt elements 1n the environment.
Scltnct, 183:1049.
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MIXING ZONES
INTRODUCTION:
A mixing zone is an area contiguous to a discharge where receiving water
quality may neither meet all quality criteria nor requirements otherwise
applicable to the receiving water. It is obvious that any time an effluent Is
added to a receiving waterway, where the effluent is poorer in quality, there
will be a zone of mixing. The mixing zone should be considered as a place
where wastes and water mix and not as a place where effluents 'are treated.
RATIONALE:
Because damage to the aquatic resource can occur when quality standards
are violated, the permissible size of a mixing zone is dependent upon the
acceptable amount of damage. The permissible size depends in part on the size
of the particular receiving water; the larger the water body, the larger the
mixing zone may be without violating quality standards in more than a given
percentage of the total area or volume of the receiving water. Likewise, the
greater number of mixing zones within a reach of river or within a water body,
the smaller each must be in order to maintain an appropriate mixing zone to
water body ratio. Future industrial and population growths must be considered
in designating such areas for wastes admixture.
As a guideline, the quality for life within a mixing zone should
be such that the 96-hour 1C far biota significant to the indigenous
50
aquatic conmunity is not exceeded; the mixing zone should be free from
effluent substances that will settle to form objectionable deposits,
free from effluent-associated materials that float to form unsightly
masses, and free from effluent-associated substances that produce object-
ionable color, odor, or turbidity. . . _
193
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A prime purpose in designating the location, size, and area constraints
of a mixing zone is to protect the aquatic life within the receiving waterway.
Shallow water areas, generally, are the nursery areas for aquatic ecosystems.
Designating offshore mixing areas or providing a larger available volume or
area for mixing offshore as a viable alternative to a smaller shoreside area
has a lesser potential for adverse biotic effects than a comparable discharge
area in shallow water. Offshore, diffusion will tend to occur in all directions
and not be constrained by a land barrier. Mixing zones may be less harmful bio-
logically when located deep within the receiving water and, wherever possible,
beneath the light-penetration area where photosynthesis occurs and algae and
associated protozoa and other organisms provide the extensive base for the
aquatic food web.
An axiom of environmental quality is that different areas vary in ecological
importance, one from the other. Generally the highest importance, and therefore
the greatest protection, must be placed on shallow-water shoreline areas of
rivers, lakes and coastal zones and on the nation's wetlands. These are
commonly the areas that protect the young and supply the food not only for the
animals that live in open waters but also for those animals that depend upon
water in some measure for their existence. Likewise, one local aquatic area
may have a higher social or ecological value than another, and the higher that
value the greater the protection from degradation that is warranted within a
waste mixing area.
Mixing zones should be located in such a manner that they do not form a
barrier to the migratory routes of aquatic species. On a given reach of a stream
or river , it would be good practice to limit the total mixing zone area to
one-third of the receiving water width. In the same fashion, the combined
m
-------
areas of all mixing zones within a lake should not exceed ten percent of
the lake surface area. In some cases, this maximum Ihould be reduced depending
on lake volume and other local conditions. Within an estuary, the maximal
dimension of the mixing area should not exceed 10 percent of the cross-
sectional area of the waterway. It is not the objective of this
rationale to outline limits for effluents, but to provide the reader
with seme of the general biological and physical considerations
necessary for the establishment of mixing zones.
In essence, the positioning of mixing zones should be accomplished in a
manner that will provide the greatest protection to aquatic life and for the
various uses of water. Generally., shoreline and surface areas for waste
admixture should be discouraged in preference to deep water, offshore designa-
tions. The relative social and ecological values of the aquatic life that may
inhabit a particular waterway area should be given due consideration in zone
definition. (Fetterolf, 1973; NAS, 1974) The designation of particular mixing
zones is a task that should follow the biological, physical and chemical
appraisal of the receiving waterway.
-------
References cited.:
Fetterolf, C.M., Jr., 1973. Mixing zone concepts; Biological Methods for the
Assessment of Water Quality, ASTM STP-528, American Society for Testing
and Materials, pp. 31-45.
National Academy of Sciences, National Academy of Engineering, 1974. Water
Quality Criteria, 1972. U. S. Government'Printing Office, Washington, D.C.
-------
NICKEL
CRITERION;
O.pl0f the 96-hour LC50 for freshwater and marine aquatic life.
"5
INTRODUCTION:
Nickel is a silver-white,metallic element seldom occurring
in nature in the elemental form. Nickel salts are soluble and
can occur as a leachate from nickel-bearing ores. Kopp and
Kroner (1967) detected nickel in the Lake Erie Basin at a frequency
of 53 percent and a mean concentration of 56 ug/1. At several
selected stations, dissolved nickel ranged from 3 to 86 ug/1 and
suspended nickel from 5 to 900 ug/1. Nickel is present in sea
water at 5 to 7 ug/1 (NAS, 1974).
RATIONALE;
Nickel is considered to be relatively non-toxic to man
(Schroeder, et al., 1961) and a limit for nickel is not
included in the EPA National Interim Primary Drinking Water Regulations
((40 FR 59566, Deoenfoer 24, 1975)lhe toxicity
of nickel to aquatic life, as reported by McKee and Wolf (1963),
indicates tolerances that vary widely and that are influenced
by species, pll, synergistic effects and other factors. The survival
curves for sticklebacks in soft tap water indicate a lethal limit
of 800 ug/1 nickel (Jones, 1939). Pickering and Henderson (1964)
reported that 96-hour LC5Q values of nickel in soft water for four
species of fish varied from 4.6 to 9.8 mg/1, and in hard
-------
water for two fish species, the 96-hour LC5Q varied from 39.2 to
42.4 mg/1. The 96-hour LCgg values for two species of aquatic
insects were reported by Warnich and Bell (1969) to be 4.0 and
33.5 mg/1 nickel. Diesinger and Christenson (1972) found that
the three-week LC50 value for Daphnia magna in soft water was 130 ug/1
nickel; 95 ug/1 .caused a 50 percent impairment in reproductivity,
and 30 ug/1 caused a 16 percent impairment.
In continuous bioassay tests designed to determine chronic
effects, Pickering (1974) demonstrated that nickel concentrations
of 380 ug/1 and lower in hard water did not adversely affect survival,
growth, or reproduction of the fathead minnow. Nickel concentrations
of 730 ug/1 caused a significant reduction both in the number of
eggs per spawning and in the hatchability of the eggs.
Calabrese, e_t al.,(1973) reported a 48-hour LC50 of 1,180
ug/1 for American oyster embryo, Crassostrea virqinica.
for larvae of the hard shell clam, Mercenaria mercenaria
(Calabrese & Nelson, 1974). Jones (1939) reported a 96-hour
LC50 of 800 ug/1 for the euryhaline stickleback, Gasterosteus
aculeatus. Gentile (1975) found that the 96-hour LC50 for the
marine copepod, Acartia tonsa was 625 ug/1.
Nickel salts have been shov/n to be injurious to plants. In
sand and nickel solution experinonts, Vanselow (1966) demonstra-
ted that at 0.5 to 1.0 mg/1, nickel is toxic to a number of
plants. The toxicity exhibited to plants by nickel varied widely
with the species. IlcKee and Wolf (1963) indicated that nickel
was extremely toxic to citrus. Chang and Shernan (1953) found
-------
that tomato seedlings were injured by 0.5 mg/1 nickel. Hop plants
were shown to be injured by nickel at 1.0 mg/1 (Legg and Ormerod,
1958). Plants exhibiting less susceptibility to nickel were:
oats, with toxic effects at 2.5 mg/1 (Orooke, 1954); corn at 2
mg/1; and tobacco with no toxic effects at 3.0 rag/1 (Soane and
Saunders, 1959).
Data indicate that, (1) nickel in water is toxic to plant
life at concentrations as low as 500 ug/1, (2) nickel adversely
affects reproduction of a freshwater crustacean at concentrations
as low as 95 ug/1, (3) marine clam larvae can be killed by
concentrations of nickel as low as 310 ug/1, and (4) reproduction
of the fathead minnow is detrimentally affected by nickel at
concentrations as low as 730 ug/1. Concentrations of nickel at
or below 100 ug/1 should not be harmful to irrigated plants or
marine and ftmfeMafer aquatic organisms.
'96
-------
REFERENCES CITED:
Biesinger, K.E. and G.M. Christensen, 1972. Effects of various njetals
on survival, growth, reproduction and metabolism of Daphnia magna.
Jour. Fish. Res. Bd. Can., 29: 1691.
Calabrese&, A., £tal_., 1973. The toxicity of heavy metals to embryos of the
American oyster, Crassostrea virginica. Marine Biology, 18:162.
Calabrese; A. and D.A. Nelson, 1974. Inhibition of embryonic
development of the hard clam, Mercenaria mercenaria.
by heavy metals. Bull. Environ. Contam. Tox. 11:92
Chang, A.T. and 6.D. Sherman, 1953. The nickel content of some Hawaiian
soils and plants and the relation of nickel to plant growth. Hawaiian
Agr. Exp. Sta. Bull., 9: 3.
Crooke, W.M., 1954. Effect of nickel versenate on oat plants. Nature
173: 403.
Gentile, J. 1975. Environmental Research Laboratory,
Narragansett, R.I., Semi-Annual Report.
Jones, J.R.E., 1939. The relation between the electrolytic
solution pressures of the metals and their toxicity
to stickleback. Jour. Exper. Blol. 16:425.
-------
Kopp, J.F. and R.C. Kroner, 1967. Trace metals in waters of the United
States. U.S. Department of Interior, FWQA, Cincinnati, Ohio.
Legg, J.T. and P.J. Ormerod, 1958. The nickel content of hop plants
with reference to nettlehead symptoms. Annual Report of the East
Mailing Research Station (England), 45: 129.
McKee, J.E. and H.W. Wolf, 1963. Water quality criteria. California
State Water Resources Control Board, Pub. No. 3-A.
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972. U.S. Government Printing Office,
Pickering, Q.H. and C. Henderson, 1964. The acute toxicity of some
heavy metals to different species of warm water fishes. Purdue
University Ext. Serv., 117: 578.
Pickering, Q.H., 1974. Chronic toxicity of nickel to the fathead
minnow. Jour. Water Poll. Cont. Fed., 46: 760.
Schroeder, H.A., ejt al_., 1961. Abnormal trace elements in man—nickel.
Jour. Chron. Dis., 15: 51.
Soane, B.K. and D. H. Sanders, 1959. Nickel and chromium toxicity of
serpentine soils in southern Rhodesia. Soil Science, 88: 322.
Vanselow, A.P., 1966. Diagnostic criteria for plants and soils.
Univ. of Calif., Div. of Agr. Sci., Berkley, California, pp. 142-156.
Warnick, S.L. and H.L. Bell, 1969. The acute toxicity of some heavy
metals to different species of aquatic insects. Jour. Water Poll.
Cont. Fed., 41: 280.
2 oo
-------
NITRATES; NITRITES
CRITERION;
10 mg/1 nitrate nitrogen (N) for
domestic water supply (health).
INTROJUCTION;
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 is either 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 feedlot
discharges. Diffuse sourcescof nitrogen include farm-site fertilizer
and animal wastes, lawn fertilizer, leachate from waste disposal in
dumps or sanitary landfills, atmospheric fallout, nitric oxide and nitrite
201
-------
discharges from automobile exhausts and other combustion processes.
and losses from natural sources such as mineralization of soil
organic matter (NAS, 1972). Water reuse systems in some fish
hatcheries employ a nitrification process for ammonia reduction;
this may result in exposure of the hatchery fish to elevated levels
of nitrite (Russo, e^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" concentrations, nitrates are rapidly ex-
creted in the urine. High intake of nitrates constitutes a hazard
primarily to warm blooded animals under conditions that are favorable
to their 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 methemo-
globin, with consequent impairment of oxygen transport.
The reaction of nitrite with hemoglobin can be hazardous in infants
under three 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/1 nitrate nitrogen
(N) (NAS, 1974). High nitrate concentrations frequently are found in
shallow farm and rural community wells, often as the result of
inadequate protection from barnyard drainage or from septic tanks
(USPHS, 1961; Stewart, eiral., 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 2000 cases of infant methemoglobinemia
-------
have been reported in Europe and North America since 1945; 7 to 8
percent of the affected infants died (Walton, 1951; Sattelmacher, 1962).
Many infants have drunk water in which the nitrate nitrogen content
was greater than 10 mg/1 without developing methemoglobinemia. Many
public water supplies in the United States contain levels that routinely
are in excess of this amount, but only one U. S. case of infant methe-
moglobinemia associated with a public water supply has ever been
reported (Vigil, 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 manuals. Methenoglobinemia symptoms 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 (N1S, 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/1 nitrate nitrogen, this level is
the criterion for domestic water supplies. Waters with nitrite nitrogen
concentrations over 1 mg/1 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
LC values for chinook salmon, Oncorhynchus tshawytscha, were 1310
50 V "~
and 1080 mg/l nitrate nitrogen in fresh water and 990 and 900 mg/1
nitrate nitrogen in 15 o/oo saline water. For fingerling rainbow trout,
Salmo gairdneri, the respective 96-hour and 7-day LC values were
1360 and 1060 mg/1 nitrate nitrogen in fresh water, and 1050 and
900 mg/1 nitrate nitrogen in 15 o/oo saline water. Trama (1954)
reported that the 96-hour LC for bluegills, Lepomis macrochirus. at
50
20° C was 2000 mg/1 nitrate nitrogen (sodium nitrate) and 420 mg/1
£04
-------
nitrate nitrogen (potassium nitrate). Knepp and Arkin (1973) observed
that largemouth bass, Micropterus salmoides. and channel catfish,
Ictalurus punctatus, could be maintained at concentrations up to 400 mg/1
nitrate (90 mg/1 nitrate nitrogen) without significant effect upon
their growth and feeding activities.
The 96-hour and 7-day LC values for chinook salmon,
50
Oncorhynchus tshawytscha, were found to be 0. 9 and 0. 7 mg/1 nitrite
nitrogen in fresh water (Westin, 1974). Smith and Williams (1974)
tested the effects of nitrite nitrogen and observed that yearling rainbow
trout, Salmo gairdneri, suffered a 55 percent mortality after 24 hours
at 0. 55 mg/1, fingerling rainbow trout suffered a 50 percent mortality
after 24 hours of exposure at 1. 6 mg/1, and chinook salmon,
Oncorhynchus tshawytscha, suffered a 40 percent mortality within 24
hours at 0. 5 mg/1. There were no mortalities among rainbow trout
exposed to 0.15 mg/1 nitrite nitrogen for 48 hours. These data
indicate that salmonids are more sensitive to nitrite toxicity than
are other fish species, e.g., minnows, Phoxinus laevis, that suffered
a 50 percent mortality within 1. 5 hours of exposure to 2030 mg/1
nitrite nitrogen, but required 14 days of exposure for mortality to
occur at 10 mg/1 (KLingler, 1957), and carp, Cyprinus carpio, when
raised in a water reuse system, tolerated up to 1. 8 mg/1 nitrite
nitrogen (Saeki, 1965).
Gillette, et al. (1952) observed that the critical range for creek
chub, Semotilus atromaculatus, was 80 to 400 mg/1 nitrite nitrogen.
Wallen, et al. (1957) reported a 24-hour LC of 1. 6 mg/1 nitrite
50
nitrogen, and 48- and 96-hour LC values of 1. 5 mg/1 nitrite
50
-------
nitrogen for mosquitofish, Gambusia affinis. McCoy (1972) tested
the nitrite susceptibility of 13 fish species and found that logperch,
Percina caprodes, were the most sensitive species tested (mortality
at 5 mg/1 nitrite nitrogen in less than 3 hours of exposure), whereas
carp, Cyprinus carpio, and black bullheads, Ictalurus melas, survived
40 mg/1 nitrite nitrogen for a 48-hour exposure period; the common
white sucker, Catostomus commersoni, and the quillback, Carpiodes
cyprinus, survived 100 mg/1 for 48 and 36 hours, respectively.
Russo, et^aL (1974) performed flow-through nitrite bioassays in
hard water (hardness = 199 mg/1 CaCOs, alkalinity = 176 mg/1 CaCOj*
pH = 7. 9) on rainbow trout, Salmo gairdneri, of four different sizes,
and obtained 96-hour LCgQ values ranging from 0.19 to 0. 39 mg/1
nitrite nitrogen. Duplicate bioassays on 12-gram rainbow trout were
continued long enough for their toxicity curves to level off, and
asymptotic LCgQ concentrations of 0.14 and 0.15 mg/1 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/1; for the
235-gram trout, the minimum level with no mortality after 10 days
was 0. 06 mg/1.
It is concluded that: (1) levels of nitrate nitrogen at or below
90 mg/1 would have no adverse effects on warm water fish (Knepp
and Arkin, 1973); (2) nitrite nitrogen at or below 5 mg/1 should be
protective of most warm water fish (McCoy, 1972); and (3) nitrite
nitrogen at or below 0. 06 mg/1 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 cold water fish could rarely occur
in nature, restrictive criteria are not recommended.
lob
-------
REFERENCES CITED;
Gillette, L.A., et^aL» 1952. Appraisal of a chemical waste problem
by fish toxicity tests. Sewage Ind. Wastes, 24:1397.
Harmeson, R. H., e£ aL, 1971. The nitrate situation in Illinois.
Jour. Amer. Water Works Assn., 63:303.
Klingler, K., 1957. Sodium nitrite, a^slow acting fish poison.
Schweiz, Z. Hydro!. 19(2):565.
Knepp, G. L. and G. F. Arkin, 1973. Ammonia toxicity levels and
nitrate tolerance of channel catfish. The Progressive Fish-
Culturist, 35:221.
McCoy, E.F., 1972. Role of bacteria in the nitrogen cycle in lakes.
Environmental Protection Agency, Water Pollution Control Research
Series, U.S. Government Printing Office (EP 2.10:16010 EHR 03/72),
Washington, D. C.
National Academy of Sciences, 1972. Accumulation of nitrate. National
Academy of Sciences, Washington, D. C.
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972. U.S. Government Printing Office,
Washington, D. C.
Russo, R. C., ei_ aL , 1974. Acute toxicity of nitrite to rainbow trout
(Salmo gairdneri). Jour. Fish. Res. Bd. Can. ,31:1653.
-------
Russo, R. C. and R. V. Thurston, 1975. Acute toxicity of nitrite to
cutthroat trout (Salmo clarki). Fisheries Bioassay Laboratory Tech.
Report No. 75-3, Montana State University.
Saeki, A., 1965. Studies on fish culture in filtered closed-circulating
aquaria. II. On the carp culture experiments in the systems. Bull.
Jap. Soc. Sci. Fish., 31:916.
Sattelmacher, P.G., 1962. Methemoglobinemia from nitrates in drinking
water. SchrReihe. Ver. Wasser-, Boden-u. Lufthyg. No. 21, Fischer,
Stuttgart.
Smith,'C.E. and W. G. Williams, 1974. Experimental nitrite toxicity
in rainbow trout and chinook salmon. Trans. Amer. Fish. Soc., 103:389.
Stewart, B. A., et al., 1967. Nitrate and other pollutants under fields
and feedlots. Envir. Sci. Tech., 1:736.
Trama, F.B., 1954. The acute toxicity of some common salts of sodium,
potassium and calcium to the common bluegill (Lepomis macrochirus
Rafinesque). Proc. Acad. Nat. Sci., Philadelphia, Pennsylvania,
106:185.
United States Public Health Service, 1961. Groundwater contamination:
Proceedings of 1961 symposium. Tech. Rpt. W61-5, R.A. Taft
Sanitary Engineering Center, U. S. Public Health Service, Department
of Health, Education and Welfare, Cincinnati, Ohio.
Vigil, J., et^aL, 1965. Nitrates in municipal water supplies cause
methemoglobinemia in infants. Public Health Rept. * 80:1119.
OPS
-------
Wallen, I.E., e^al., 1957. Toxicity to Gambusia affinis of certain
pure chemicals in turbid waters. Sewage Ind. Wastes, 29:695.
Walton, G., 1951. Survey of literature relating to infant methemoglo-
binemia due to nitrate-contaminated water. Amer. Jour. Public
Health, 41:986.
Westin, D. T., 1974. Nitrate and nitrite toxicity to salmonid fishes.
The Progressive Fish-Culturist, 36:86.
Wblff, I.A. and Nwerman, 1972. Nitrates, nitrites, and nitroswdnes
•ciflno*, 177:15.
-------
OIL AND GREASE
CRITERIA;
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) 0.01fpf 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 (Blunier, 1970). A
major difficulty encountered in the setting of criteria for oils 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 non-volatile, soluble or insoluble,
persistent or easily degraded.
RATIONALE;
Field and laboratory evidence have demonstrated both acute lethal
toxicity and long-term sublejthal . toxicity of oils tP aquatic organisms.
Events such as the Tampico Maru wreck of 1957 inBaja, California,
(Diaz-Piferrer, 1962), and the No. 2 fuel oil spill in West Falrnouth,
210
-------
Massachusetts,in 1969 (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 2500 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 LC
50
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/1.
The long-term sublethal effects of oil pollution refer to interferences
with cellular and physiological processes such as feeding and reproduction
vahd 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/1 (see Table 7 ).
Table 7 summarizes some of the sublethal toxicities for various
petroleum pollutants and various aquatic species. In addition to sublethal
effects reported at the 10 to 100 ug/1 level, it has been shown that
petroleum products can harm aquatic life at concentrations as low as
1 ug/1 (Jacobson and Boylan, 1973).
Bioaccumulation of petroleum products presents two especially
important public health problems: (1) the tainting of edible, aquatic
-------
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Ucaie, 1967'1
Strand, et al. , 1971*1
'fJ3zi, 1973*1
Oordor and Prouse,
1973*-
ifilber, 1968*2
Brown, 1972*1
Bice, 1973
41ronor, 1967
h'ilsor., 1970«2
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Alien, 1971*2
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total C" * Fixata"
Avoidance eff ect '•> , couj'l
have < f ; ect on r.irr'ii.j *-n
behevfor
lrro;uler]i.y sr/1 1-1 iv IB
•Jcforrci* r.t.J in' '.-i\'f
,
and fce^lne *>"'r.'i-'>'r
ri'l/'-Sf effect, c.'. b".v,ii.,'Ic
delay rolt to ltv. r f M-«J
1..1- I'll hi' '•'•• "- In
-------
SV.WAW Pi- ik'Vi, s.iwrrtuu. tvyKv, OK pmauw i-to'.wrs os MAMVS- ;.ur
TYP O'kuVN^:" *.n\:i«S
F1--L Cliii.ook salror, (CV •> O';t^ tu'b'.i'*
' rnxtt V i • )
1 !',-• Mia lor.iJl.a
CRUSTACEANS lobster, (io^arus aneric?.rius)
Pollicires yolyrei us
K--,'N.-». lni-n-:S,Mv^v:n>-,,, j W-.^.M ros
i
lt-»-fc.-«n 4 Hnil^y, . i , .-
10,.,«l l'o:;r.fiie I 5, 3 0 y\v\
i i
Gar^j.er, et al,lO(3 Cruie (v*olo fr-iotiv-s j
(vat «•**-! *•-•."•". iV.c '
r'8 T-.- (v'vl
,.ll.,i,|TA. h> .,vv,...
res; .rot ion.
oho ' rreoeptcrs ,
Bluner, et «lt!973 ,CruJe, kerosene
Straucliaii, 1971 Cru5e-Sa.nta bartara
Lobster (H. anerici •. ;s)
Pachycrapsus crass iips
, to£^i^
i&LLUSCS Kussel 0-Vtilus eJuMs)
. Snail (Nassarius etsclctus)
T ' '
Snail (Kassarius obEoletus)
-
Claiu (y.ytt arenaria)
Oyster (Cr as so st re g
Stall (Littorir.& littorpa)
MusGel (f^rtllus e3ul;r )
i CTi^ER 1 P^/3ycha^ta (Crititeiiit
10 ppa
Field sxua/ after
blovcut .
Effe>-LS of clicnore-cer
behr 5pr, r -rrcssi^M.
^prLro'it decrease i\
brae-dins; r.o recruit
t*P». *- • <5**>i- T 3T9 " a ^sa Crude , Extracts ' r>el?.y in feeding.
Kittredge, 1972 «1 ' Grade j r-Uutic!»s of dio-.hyl | Inhlcition of fcedir;
Krebs, 1973fci Ko. 2 Fuel oil
GiU ill&n ,1973*2 , Crude
1
Eluraer, et al,1973 Kerosei.o
t
Jaectson & Boylan Kerosene
Barry end Yevich, ', llo. 2 fuel oil
137«
Ilackin and Horkios, ! Bleedvater
1961 l
' Perkh.s, 1970»1 IB? 1002
I
:Wscr.-<-~ith, 1973
Elu^-cr, et al, 19Y1*1, .'io., ? fuel oil
Eello.^, «t c.1, 1972* [Determent
"•'OT .
eiit ir.
after V.Fc.lnoutr soill ,• he!.-rior.
1 ppa
Reduction in carlon t
(increase in rcspirr.*.
Saturated extract J
diluted 10 10 ^OJ? reductin in C!-Q«C.'.
. ln«rcfirtton of fooi.
0.001 - 0.004 ppn
collected fron field
30 ppre
0.01 pp»
collected fro-a field
0.01-10 pt3
UjRDt
ion ;
&ctic
Reductin in chcmot&ctic •
perception of food
Gonadel tumors
Peduccd growth an-i e,l,
conttnt
significant inhiWtitr
p.rcvth
rogen
to
Karkel tainting
Ir.hltition in dcvc-lot'
cnt
U'-crca^e in survi /al ,
r'.car.-Hty
Mote: tl=tflier. froia National Acadtcsy of Scisr.cco, 1975
*?*U.ken froa Kocre, Jvy^r, and Katz. 1973-
-------
REFERENCES CITED IN TABLES 6 and 7
I Allen, H., 1971. Effects of petroleum fractions on the early development
of a sea urchin. Marine Pollution Bulletin, 2:138.
ATEMA, J. and L. Stein. 1972. Sublethal effects of crude oil on
t. the behavior of the American lobster. Technical Report
Woods Hole Oceanographic Inst., No. 72-72.
^ Aubert, M.R. eJ^aL, 1969. Etude de la toxicite de produits chemiques
vis-a-vis de la chaine bibliogique marine. Rev. Int. Oceanogr. Med.,
13/14:45.
U BAKER, ,J. M. 1971. Several papers in E. B. Cowell (ed.) The
I ecological effects of oil pollution on littoral communities.
Institute of petroleum, London. 250 p.
C BARRY, M. and P. Yevich, 1974. Incidence of cancer In the soft-
^ shell clam, Mya arenaria. Final report of State of Maine
(Dept. Sea and Shore Fisheries)' to U. S. Air Force. Contract
No. F. 33600-72-C-0540. 32p.
£ Bellan, G., e^aL . 1972. The sublethal effects of a detergent on the
reproduction, development, and settlement in the polychaetous annelid
Capitella capitata. Marine Biology, 14:183.
7 Blumer, M., et aL , 1971. A small oil spill. Environment, 13:2.
97 P.
O Brockson, R.W. and H. T. Bailey, 1973. Respiratory response of juvenile
Chinook salmon and striped bass exposed to benzene, a water-soluble
component of crude oil. In: Proceedings. Joint Conference on Preven-
tion and Control of Oil Spills, Amer. Pet. Inst., Washington, D. C.
\Q Brown, D. H. 1972. The effect of Kuwait crude oil and a solvent emul-
sifier on the metabolism of the marine lichen, Lichina pygmaea.
Marine Biology, 12:309.
I ( CAIRNS AND SCHEIER, 195-p.
^ CHIPMAN, W.- A. and P. S. Caltsoff. 1949. Effects of oil mixed
I with carbonized sand on aquatic animals. Spec. Scient
Rep. U. S. Fish Wildl. Serv. 1. 52 p.
-------
DORRIS, T. C. , W. Gould, and C. R. Jenkins. 1959. __Toxicity ___
bioassay of oil refinery effluents in Oklahoma. In
Trans. 2nd Sem. Biol. Prob. Water Poll. . R. A. Taft
San, Eng. Center, Cincinnati, Ohio, Tech. Rep. WGO-3:
276-285.
ENVIRONMENTAL PROTECTION AGENCY. 1974. Waste Oil Study. Report
to Congress. 401 p.
,." Gardner, G.R. , e^aJL , 1973. Analytical approach in the evaluation
of biological effects. Jour. Fish. Res. Bd. Canada, 35:3185.
Gilfillan, E.G., 1973. Effects of seawater exracts of crude oil on
carbon budgets in two species of mussels. In: Proceedings. Joint
Conference on Prevention and Control of Oil Spills. Arner. Pet.
Inst. , Washington, D. C.
// Gordon. D. C. and N. J. Prouse. (ialjLij. The effects of three
different oils on marine phytoplankton photosynthesis'. -Marine Biol.
f^j Jacobson, S. M. and Eoylan, 1973. Effect of seawater soluble fraction
of kerosene on chemotaxis in a marine snail, Nassarius obsoletus.
Nature. 241:213.
la KARINEN, J. F. and S. D. Rice. • 1974. Effects of T?rudhoe Bay crude
7 oil on molting tanner crabs, Chionoecetes bairdi. Marine
Fisheries Review. 36:31.
2j£ Kauss. el^al; . 1972. Field and laboratory studies of the effects of crude
oil spills on phytoplankton. In: Proceedings, 18th Annual Technical
Conference, Environmental Progress in Science and Education.
o i Kittredge, J. S. , 1973. Effects of water-soluble component of oil
pollution on chemoreception by crabs. Fisheries Bulletin.
T, > Krebs, C.T., 1973. Qualitative observations of the marsh fiddler
(Uca Pugnax) populations in Wild Harbor Marsh following the
September, 1969 oil spill. National Academy of Sciences. Washington.
D.C., Unpublished manuscript.
-------
Kuhnhold, W.W., 1970. The influence of crude oils on fish fry. In;
Proceedings, FAO conference, Rome, Italy.
Lacaze, J. C.. 1967. Etude de la croissance d'une algue
planctonique en presence d'un detergent utilize pour la
destruction des nappes de petrole en mer. C.R. Acad.
Sci. (Paris) 265 (Ser. D):489.
MACKIN, J. M. and S. H. Hopkins. 1961. Studies on oyster mortality
in relation to natural environments and to oil fields in
Louisiana. Publs. Inst. Mar. Sci. Univ. Tex. 7:1-131.
MEINCK, F. jet^ al._ 1956. Industrie-Abwasser. 2nd Edit. (Gustov
Fischer Verlag, Stuttgart), p. 536, 48 D. M.
Mironov. O. G.. 1967. Effects of low concentrations of petroleum
and its products on the development of r.oe of the Black Sea
flatfish. Vop. Ikhtiol. 7(3):557.
Mironov. O. G., 1970. The effect of oil pollution on flora and
fauna of the Black Sea. &: Proceedings. FAO Conference on
Marine Pollution and its Effects on Living Resources and Fish.
Rome. December, 1970. Food and Agriculture Organization
of the United Nations. Rome.
MOORE S F. R. L. Dwyer and S. N. Katz. 1973. A preliminary
•*°° ' assessment oflhe environmental vulnerability of Machxas
Bay, Maine to oil supertankers. Report No. MITSG 73-6. 162 p.
MORROW J. E. 1974. Effects of crude oil and some of its£c°mP°^n"
on young coho and sockeye salaon. Publication EPA-660/3-73-018.
U. S. E. P. A.
/2 / NELSON-SMITH, A. 1973. Oil pollution and marine ecology. Plenum
* ' press. New York. 260 p.
-7 •) Nuzzi. R., 1973. Effects of water soluble extracts of oil on phyto-
plankton. In: Proceedings, Joint Conference on Prevention and
Control of Oil Spills. Amer. Pet. Inst.. Washington, D. C.
217
-------
Perkins. E. J.. 1970. Some effects of "detergents" in the marine
environment. Chem. Ind.. 1:14.
PICKERING, Q. H. and C. Henderson. 1966. Acute toxlcity of some
important petrochemicals to fish. Water Poll. Contr.- Fed.
J. 38 (9);1419-1429.
Rice, S. D.. 1973. Toxicity and avoidance tests with Prudhoe Bay oil
and pink salmon fry. Proceedings of joint conference on prevention
of oil spills, Wash.. D. C. , pp. 667-670.
• i
Steel, D.L. and B.J. Copeland, 1967. Metabolic responses
of some estuarine organisms to an industrial effluent
control. Mar. Sci. Univ. Texas 12:143-159
Strand, J. W., et_al., 1971. Development of toxicity test procedure
for marine phytoplankton, p. 279-286. In_: American Petroleum
Institute. Proceedings of a joint conference o'n prevention and control
of oil spills, Washington, D. C.
the
?«
Calif. 426 p.
3 *T Ted, J.H. 1972. An introduction to environmental ethology
Woods Hole Oceanographic Institution. Ref. 72-42. Woods
Holei Mass. Unpublished manuscript.
Vaughan, B.E. 1973. Effects of oil and chemically disparsed
oil on selected marine biota - a laboratory study.
Richland, Washington, Battelle Pacific Northwest Laboratories
120 p.
/ Wells. P. G., 1972. Influence of Venezuelan crude oil on lobster
larvae. Marine Pollution Bull., 3:105,
218
-------
Wilber, C.G., 1968. Biological aspects of water pollution.
.^ Wilson, K. W., 1970. The toxicity of oil-spill dispersants to the
*/
embryos and larvae of some marine fish. In; Proceedings,
FAO Conference, Rome, Italy.
Wohlschlog, D.E. and J.N. Cameron. 1967. Assessment of low level,,
Stress OR the ^espirdtory metabolism of trie pinfish (uayodpn
rhomboides). Inst. Mar. Sci. Univ. Texas 12:160-171
2I&.I
-------
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/1 of crude oil caused
tainting in oysters. Moore, et^aL (1973) reported that concentrations
as low as 1 to 10 ug/1 could lead to tainting within very short periods
of time. It has been shown that chemicals responsible for cancer
in animals and man (such as 3,4-benzopyrene) occur in crude oil
(Blumer, 1970). It has also been shown that marine organisms "are
capable of incorporating potentially carcinogenic compounds into their
body fat where the compounds remain unchanged (Blumer, 1970).
Oil pollutants may also be incorporated into sediments. There is
evidence that once this occurs in the sediments below the aerobic surface
layer, petroleum oil can remain unchanged and toxic for long periods, since its rate
of bacterial degradation is slow. For example, Blumer (1970) reported
that No. 2 fuel oil incorporated into the sediments after the West Falmouth
spill persisted for-over a year, and even began spreading in the form of
oil-laden sediments to more distant areas that had remained unpolluted
immediately after the spill. The persistence of unweathered oil within
the sediment could have a long-term effect on the structure of the benthic
community or cause the demise of specific sensitive important species.
Moore, et^al. (1973) reported concentrations of 5 mg/1 for the carcinogen,
3,4-benzopyrene in marine sediments.
2«9
-------
Mironov (1967) reported that 0.01 mg/1 oil produced deformed and
inactive flatfish larvae. Mironov (1970) also reported 'ijihUiition or
delay of cellular division in algae by oil concentrations of 10~4 to
10'1 mg/1. Jacobsen 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/1. Bellen, et al. (1972)
reported decreased survival and fecundity in worms at concentrations
of 0.01 to 10 mg/1 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 petro-
chemical components.
There is a paucity of toxicological data on the ingestion of the
components of refinery wastewaters by humans or by test animals. It
is apparent that any tolerable health concentrations for petroleum derived
substances far exceed the limits of taste and odor. Since petroleum
derivatives become organoleptically objectionable at concentrations far
below the human chronic toxicity, it appears that hazards to humans will
not arise from drinking oil-polluted waters (Johns Hopkins University,
1956; Mckee and Wolf, 1963). Oils of animal or vegetable origin generally
are non-toxic 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 due to lack of mobility, (b) lethal effects on fish by coating
epithelial surfaces of gills, thus preventing respiration, (c) potential fish
kills due to 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 shore-
lines 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."
2.21
-------
Oils of animal or vegetable origin generally are chemically non-toxic to
humans or aquatic life; however, floating sheens of such oils result in
deleterious environmental effects described in this criterion.
Thus, it is recommended that surface waters shall be virtually free from
floating non-petroleum oils of vegetable or animal origin. This same
recommendation applies to floating oils of petroleum origin since they too
may produce the above effects.
221-1
-------
REFERENCES CITED;
Bel Ian, et aL, 1972. The sublethal effects of a detergent on the
reproduction, development, and settlement in the polychaetous
annelid Capitella capitata. Marine Biology, 14:183.
Blumer, M., 1970. Oil contamination and the living resources of
the sea. F.A.O. Tech. Conf. Rome. FIR:MP/70/R-1,11P.
Bugbee, S. L. and C. M. Walter, 1973. The response of macro-
invertebrates to gasoline pollution in a mountain stream. In:
Prevention and control of oil spills, proceedings of symposium
March 13-17, Washington, D. C., p. 725.
Diaz-Piferrer, 1962. The effects of an oil spill on the shore of
Guanica, Puerto Rico (abstract) Ass. Isl. Mar. Labs, 4th Mtg.
Curacao, 12-13.
Hampson, G. R. and H. L. Sanders, 1969. Local oil spill.
Oceanus, 15:8.
Jacobson, S. M. and D. B. Boylan, 1973. Effect of seawater
soluble fraction of kerosene on chemotaxis in a marine snail,
Nassarius obsoletus. Nature, 241:213.
Johns Hopkins University, 1956. Final report to the water quality
subcommittee of the American Petroleum Institute, Project
PG 49.41.
McKee and Wolf, 1963. Water quality criteria. California State
Water Resources Control Board, Pub. No. 3-A.
-------
Mlronov, O.G., 1967. Effects of low concentrations of petrol** and
its products on the development of roe of the Black Sea flatfish.
Vop Ikhtiol., 7:557.
Mironov, O. G., 1970. The effect of oil pollution on flora and
fauna of the Black Sea. In Proceedings: FAO Conference on
Marine Pollution and its effects on living resources and fish.
Rome, Dec., 1970, E-92. Food and Agriculture Organization
of the United Nations.
Moore, S. F., e£aU 1973. A preliminary assessment of the
environmental vulnerability of Machias Bay, Maine to oil
supertankers.
Nelson-Smith, A., 1971. Effects of oil on marine plants and animals,
p. 273-380. In.: P. Hepple (ed.), Water Pollution by Oil.
I.P. London.
-------
DISSOLVED OXYGEN
CRITERIA:
Aesthetics: Water should contain sufficient dissolved oxygen
to maintain aerobic conditions in the water column and, except as
affected Toy natural phenomena, at the sediment-water interface.
Freshwater aquatic life: A minimum concentration of dissolved
oxygen to maintain good fish populations is 5.0 mg/liter. The
criterion for salmonid spawning beds is a minimum of 5.0
mg/liter in the interstitial water of the gravel.
INTRODUCTION;
Dissolved oxygen historically has been a major constituent of
interest in water quality investigations. It generally has been
considered as significant in the protection of aesthetic qualities of
water as well as for the maintenance of fish and other aquatic life.
Traditionally, the design of waste treatment requirements was based
on the removal of oxygen demanding materials so as to maintain the
dissolved oxygen concentration in receiving waters at prescribed levels.
Sophisticated techniques have been developed to predict the dissolved
oxygen concentration under various hydrologic, hydrographic, and waste
loading conditions (Velz, 1970). Dissolved oxygen concentrations are
an important gage of existing water quality and the ability of a water
body to support a well balanced aquatic fauna.
RATIONALE;
The aesthetic qualities of water require sufficient dissolved oxygen
present to avoid the onset of septic conditions with its attendant
malodorous emissions. Insufficient dissolved oxygen in the water column
causes the anaerobic decomposition of any organic materials present.
Such decomposition tends to cause the formation of noxious gases such
as hydrogen sulfide and the development of carbon dioxide and methane
in the sediments which bubble to the surface or which tend to float
settled sludge as mats which are composed of various organic materials.
Dissolved oxygen in bodies of water used for municipal water supplies
Is desirable as an indicator of satisfactory water quality in. terras of low
residuals of biologically available organic materials. In addition,
dissolved oxygen in the water column prevents the chemical reduction and
subsequent leaching of iron and manganese principally from the sediments
(Environmental Protection Agency, 1973). These metals cause additional
expense in the treatment of water or affect consumers1 welfare by causinc taste
and staining plumbing fixtures and other surfaces which contact the
water in the presence of oxygen (NAS, 1974).
-------
Dissolved oxygen also is required for the biochemical oxidation of
ammonia ultimately to nitrate in natural waters. This reduction of
ammonia reduces the chlorine demand of waters and increases the
disinfection efficiency of chlorination (NAS, 1974).
The disadvantage of substantial quantities of dissolved oxygen in
water used as a source of municipal water supply is the increased rates
of corrosion of metal surfaces in both the water treatment facilities
and in the distribution system (NAS, 1974).
Such corrosion, in addition to the direct damage, can increase the
concentration of iron (and other metals) which may cause taste in the
water, as well as staining.
A discussion of oxygen criteria for freshwater fish must take into
account these facts: (1) fish vary in their oxygen requirements according
to species, age, activity, temperature, and nutritional state; (2) they
are found from time to time, and can survive for a while,
at oxygen concentrations considerably below that considered suitable for
a thriving population; and (3) although there is much literature on the
oxygen consumption of fish and the effects of varying oxygen concentrations
on behavior and survival, few investigators have employed methods or
sought endpoints that can be related with confidence to maintaining a
good fish population.
To allow for the differences among requirements affected by species
and other variables, the dissolved oxygen criteria are based on the
concentration that will support a well-rounded population of fish (Ellis,
1937) as it would occur under natural conditions. A population of fish
is composed of a number of different but more or less interdependent
species, of different feeding and reproductive habits, but which wilj.
include game and pan fish (bass, pike, trout, perch, sunfish, crappie,
depending upon the location), some so-called rough or coarse fish
(carp, buffalo, bullhead, sucker, chub), and large numbers of smaller
'forage* fish (e.g., minnows). Theoretically it should be possible
to base oxygen criteria on the needs of the most sensitive component
of such a population, but there is not enough information for this at
present; that is why the criteria must be based on oxygen concentrations
known to permit the maintenance and well-being of the population as a
whole.
The requirement that the data be applicable to naturally occurring
populations imposes limits on the types of research that can be used as
a basis for the criterion. Aside from a few papers on feeding, growth,
and survival in relation to oxygen concentration, very little of the
laboratory-based literature has a direct bearing; field data are in
general more useful. Field studies have the disadvantage that the
numbers of variables encountered in the natural environment (temperature,
pH, dissolved solids, food supply, and the like, as well as dissolved
oxygen) make it necessary to be conservative in relating fish abundance
and distribution to oxygen concentration alone, but enough observations
have been made under a variety of conditions that the importance of
oxygen concentration seems clear.
-------
Field studies, in which fish catches, have been related to dissolved
oxygen concentrations measured at the same time, indicate that a dissolved
oxygen concentration of 3 mg/liter is too low to maintain a good fish
population (Thompson, 1925; Ellis, 1937; Brinley, 1944), and this
finding is supported by laboratory observations that in the vicinity of
3 mg/liter and below feeding .is diminished or stopped (Lindroth, 1949;
Mount, 1960; Herrmann, et^ al_., 1962), and growth is reduced
(Hamdorf, 1961; Itazawa, 1971), even when the lowered oxygen concentration
occurs for only part of the day (Stewart, et al._». 1967) .
A dissolved oxygen concentration of 4 mg/liter seems to be about
the lowest that will support a varied fish population (Ellis, 1937),
even in the winter (Thompson, 1925), and for a well-rounded population
including game fish it should be above that. Both Ellis (1937) and
Brinley (1944) set the minimum for a well rounded population at 5 mg/liter.
It should be pointed out, however, that Thompson found the greatest
variety of species at 9 mg/liter, Ellis found good populations more
frequently at 6 than at 5 mg/liter, and Brinley reported the best
concentrations for game fish populations to be above 5 mg/liter. The
belief that 5 rag/liter is adequate is supported by the fact that the
introduced rainbow trout thrives in Lake Titicaca (Everett, 1973) where,
because of the altitude, the oxygen concentration in fully saturated
water is not over 5 mg/liter.
Fish embryonic and larval stages are especially vulnerable to reduced
oxygen concentrations because their ability to extract oxygen from the
water is not fully developed and they cannot move away from adverse
conditions. Although many species can develop at oxygen concentrations
as low as 2.5 to 3 mg/liter, the effects of a reduced oxygen concentration
even as high as 5 or 6 mg/liter can cause a partial mortality or at the
least retard development (Brungs, 1971; Siefert e± ail., 1973, 1974, 1975;
Carlson et al., 1974; Carlson and Siefert, 1974; Garside, 1966; Gulidov,
1969; Hamdorf, 1961). Unless it is extreme, however, the retardation
need not be permanent or detrimental to the species (Brannon, 1965;
Eddy, 1972). For most fish, maintaining a minimum of 5 mg/liter in the
water mass in the vicinity of the embryos and larvae should suffice.
Special treatment is required for species, such as the salmonids,
that bury their fertilized eggs in gravel. The flow through gravel is
often slow, especially if siltation has occurred, and if it is slow
enough the developing fish and other organisms can easily deplete the
oxygen supply enough to cause damage, especially if the concentration in
the water is relatively low before it enters the gravel (Cooper, 1965;
Coble, 1961; Brannon, 1965). With a permeable gravel and abundant flow,
5 mg/liter in the overlying water should be enough. This concentration
could well be inadequate, however, with a less porous gravel and a slower
flow. Since the permeability and flow have so important a bearing on
the initial oxygen concentration required to maintain the intragravel
concentration, and since these characteristics vary with location, it is
proposed that the criterion for salmonid spawning beds be stated as not
less than 5 mg/liter in the gravel. This would require that the concentration
in the water entering the gravel be 5 mg/liter or more, increasing as the
intragravel flow rate decreased.
-------
4
Decreased dissolved oxygen levels, if sufficiently severe, can
adversely affect aquatic insects and other animals upon which fish feed.
Sprague (1963) has evaluated such effects on several crustaceans while
others have evaluated caddisfly larvae and stonefly nymphs (Doudoroff
and Shumway, 1970). However, many other invertebrates are less sensitive
to lowered dissolved oxygen concentrations and may be equally suitable
fish food. Doudoroff and Shumway (1970) concluded that as long as
dissolved oxygen concentrations remain entirely satisfactory for fish,
no material impairment of the food resources for fish ascrihable
to dissolved oxygen insufficiency will occur.
227
-------
LITERATURE CITED:
Brannon, E. L. 1965. The influence of physical factors on the development
and weight of sockeye salmon embryos and alevins. International
Pacific Salmon Fisheries Commission, Progress Report No. 12, pp. 1-26.
mimeo.
Brinley, F. J. 1944. Biological studies. House Document 266, 78th
Congress, 1st session; Part II, Supplement F. pp. 1275-1353.
Brungs, W. A. 1971. Chronic effects of low dissolved oxygen concentrations
on fathead minnow (Pimephales promelas). J. Fish. Res. Bd. Canada.
28: 1119-1123.
Carlson, A. R., et al., 1974. Effects of
lowered dissolved oxygen concentrations on channel catfish (Ictalurus
punctatus) embryo and larvae. Trans. Amer. Fish. Soc. 103: 623-626.
Carlson, A. R. and R. E. Siefert. 1974. Effects of reduced oxygen on the
embryos and larvae of lake trout (Salvelinus namaycush) and largemouth
bass (Micropterus salmoides). J. Fish. Res. Bd. Canada. 31: 1393-1396.
Coble, D. W. 1961. Influence of water exchange and dissolved oxygen in
redds on survival of steelhead trout embryos. Trans. Amer. Fish. Soc.
90: 469-474.
Cooper, A. C. 1965. The effect of transported stream sediments on the
survival of sockeye and pink salmon eggs and alevin. International
Pacific Salmon Fisheries Commission, Bulletin 18: 1-71.
Doudoroff, P. and D. L. Shumway. 1970. Dissolved oxygen requirements of
fresh water fishes. FAO Fish. Tech. Paper No. 86.
Eddy, R. M. 1972. The influence of dissolved oxygen concentration and
temperature on the survival and growth of chinook salmon embryos and
fry. M.S. Thesis, Oregon State Univ., June 1972.
Ellis, M. M. 1937. Detection and measurement of stream pollution. Bull.
U.S. Bureau of Sport Fisheries, and Wildlife. 48(22): 365-437.
Environmental Protection Agency. 1973. The control of pollution from
hydrographic modifications. EPA 430/9-73-017, U.S. Government
Printing Office, Washington, D.C.
Everett, G. V. 1973. Rainbow trout, Salmo gairdneri (Rich.), fishery of
Lake Titicaca. J. Fish. Biol. 5: 429-440.
Garside, E. T. 1966. Effect of oxygen in relation to temperature on the
development of embryos of brook trout and rainbow trout. J. Fish.
Res. Bd. Canada. 23: 1121-1134.
228
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Gulidov, M. V. 1969. Embryonic development of the pike (Esox lucius L.)
when incubated under different oxygen conditions. Prqbs. of Ichthyol.
9: 841-851.
Hamdorf , K. 1961. Die Beeinflussung der Embryonal- und Larvalentqicklung
der Regenbogenforelle (Salmo irideus Gibb.) durch die Umwelgf actor en
02-Partialdruck und Temperatur. Z. vergl. Physiol. 44: ^51-462.
Herrmann, R. B., et al. , 1962. Influence of
dissolved oxygen concentrations on the growth of juvenile coho salmon.
Trans. Amer. Fish. Soc. 91: 155-167.
Itazawa, Y. 1971. An estimation of the minimum level of dissolved oxygen
in water required for normal life of fish. Bull. Jap. Soc. Sci. Fish.
37: 273-276.
Lindroth, A. 1949. Vitality of salmon parr at low oxygen pressure. Inst.
Freshwater Res. Drottningholm, Report #29. Fish. Bd. of Sweden (Annual
report for 1948): 49-50.
Mount, D. I. 1960. Effects of various dissolved oxygen levels on fish
activity. Ohio State Univ. Natur. Resources Inst., Ann. Fisheries
Res. Rept. pp. 13-33.
National Academy of Sciences/ National Academy of Engineering. 1974.
Water quality criteria, 1972. U.S. Government Printing Office,
Washington, D.C.
Siefert, R. E., et al . , 1973. Effects of reduced
oxygen concentrations on northern pike (Esox lucius) embryos and larvae.
J. Fish. Res. Bd. Canada. 30: 849-852.
Siefert, R. E. and W. A. Spoor. 1974. Effects of reduced oxygen concentrations
on embryos and larvae of white sucker, coho salmon, brook trout, and
walleye. Proceedings of an International Symposium on the Early Life
History of Fish. Oban, Scotland, May 17-23, 1973, Edited by J. H. S.
Blaxter, Springer-Verlag Berlin Heidelberg New York. pp. 487-495.
Siefert, R. E., et al . , 1975. Effects of reduced
oxygen concentrations on the early life stages of mountain whitefish,
smallmouth bass, and white bass. Accepted for publication in the Prog.
Fish. Cult.
Sprague, J. B. 1963. Resistance of four fresh water crustaceans to lethal
high temperature and low oxygen. J. Fish. Res. Bd. Canada. 20: 387,
Stewart, N. E., *-&!, . . %. 1967. Influence of oxygen
concentration on the growth of juvenile largemouth bass. J. Fish. Res.
Bd. Canada. 24: 475-494.
Thompson, D. H« 1925. Some observations on the oxygen requirements of
fishes in the Illinois River. 111. Nat. Hist. Surv. Bull. 15: 423-437.
Velz, C. J. 1970. Applied stream sanitation. Wiley-Interscience, New York.
22?
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AIDRIN-DIELJ3RIN
CRITERIA;
.003 ug/1 for freshwater and marine aquatic life.
The persistence, bioaccunulaticn potential
and carcinogenicity of aldrin-dieldrin cautions
human exposure to a minimum.
RATIONALE;
Since dieldrin is a highly persistent chemical which bioacxumulates
in aquatic organisms used for human food and is also considered a potential
I
human carcinogen, levels of dieldrin in waterways should be kept as low
as feasible. Action by the Environmental Protection Agency in suspending
the production and use of dieldrin should result in a gradual decrease
in concentrations in the environment. Such
additions should not be permitted without substantial documentation that
alternatives are either infeasible or potentially more hazardous. The
persistence, bioaccumulative properties and carcinogenic potential of
dieldrin should be taken into account when determining the uses of water
with measurable amounts of dieldrin.
Aldrin is metabolically converted to dieldrin by aquatic organisms.
Residues in goldfish, Carassius auratus, exposed to aldrin for 32 days
were found to consist of 93.9 percent or more dieldrin except for
visceral fat where residues were 100 percent dieldrin after 31.5 to
92.4 days' exposure (Gakstatter, 1968). Epoxidation of aldrin to dieldrin
was found to occur also at lower trophic levels. The relative rates of
this conversion were low for algae and high for the protozoa (Kahn, et al.,
1972). Because of this metabolic conversion and because of evidence that
dieldrin is as toxic or slightly more toxic than aldrin to aquatic organisms
236
-------
(Jensen and Gaufin, 1966; Henderson, et al_., 1959), an acceptable
water concentration is based on the presence .pf either aldrin or dieldrin
or the sum of both.
In toro studies comprising five long-term oral studies feeding dieldrin
to CF-1 mice at various concentrations, liver enlargements and tumors were
detectable. Appearance of tumors was dose responsive since tumors occurred
9 months following treatment with 10 ppm; 19 months with 5 ppm; and 23
months with 2.5 ppm. Further, the group all experienced a decrease in
survival rates. At intake rates of 1.25 ppm and 1 ppm dieldrin, no liver
enlargements were detected clinically and survival was not affected (Walker,
et al., 1972).
The best evidence that aldrin-dieldrin poses a cancer hazard to man is
provided by mouse laboratory data. Although the liver was the principal
organ affected in the mouse and was a major site of action in rats, there
was an increase of tumors in the lungs and other organs (Keller, et al., 1974).
Ninety-six-hour LC50 values of 16, 7.9, and 8.5 ug/1 dieldrin have
been reported for the fathead minnow, Pijnephales promelas; bluegill,
Lepomis macrochirus; and green sunfish, Lepomis cyanellus, respectively
(Tarzwell and Henderson, 1956).
The LC50 for the stonefly naiad, Acroneuria pacifica, exposed to
dieldrin in a continuous flow bioassay system for 20 days was 0.2 ug/1
(Jensen and Gaufin, 1966). The 48-hour BC50 (inmobilization value at 15°C) for
daphnids, Siirocephalus serrulatus and Daphnia pulex, to dieldrin was
0.24 mg/1 and 0.25 mg/lf respectively (Sanders and Cope, 1966).
23>
-------
The sailfin molly, Poecilia latipinna/ appears to be the roost
sensitive freshwater fish species tested for chronic effectr; growth
rates and reproductive performance were adversely affected during a
34-week exposure to dieldrin at 0.75 ug/1 (Lane and Livingston, 1970).
Guppy, Poecilia retlculata, populations were affected by levels of
dieldrin at 1.8 ug/1 water concentration during a 14-month exposure.
Exposed populations developed greater total numbers of individuals than
did controls. This phenomenon may have been caused by slight, change in
the feeding behavior of the adults induced by dieldrin, consisting of
inhibition of the normal predation by the adults upon the fry (Cairns,
et al., 1967).
Residue accumulation of dieldrin and aldrin is well documented. Levels
of dieldrin in fish tissue from Lake Michigan have been as much as
100,000 times (wet weight basis) the dieldrin levels occurring in the
water (Reinert, 1970). Lake water concentrations in Lake Michigan were
in the 1 to 3 nanogram/liter range, and whole-fish concentrations ranged
from a low of 0.03 ppm for lake whitefish to 0.20 ppm for lake herring,
0.23 ppm for bloater, and 0.28 ppm for kiyi. Laboratory exposures of
fish, invertebrates, and algae have indicated that residue accumulation
232
-------
of aldrin and dieldrin is significant. The Reticulate sculpin,
Cottus perplexus, exposed to 8.6, 1.7, 0.86, 0.17, 0.086, and 0.017 ug/1
dieldrin in water for 32 days were found to have tissue concentrations
(wet weight basis) often exceeding 50,000 times the water exposure
level (Chadwick and Brockson, 1969). The sailfin molly, Poecilia
latipinna, exposed for 34 weeks, at 12, 6, 3, 1.5, and 0.75 ug/1 dieldrin
in water concentrated dieldrin in all tissues (wet weight basis) at
least 10,000 times (Lane and Livingston, 1970). At the termination
of a 64-week exposure of the ostraced, Chlamydotheca arcuata, to water
concentrations of aldrin at 0.01 and 0.10 ug/1 and dieldrin at 0.01 and
0.10 ug/1, dieldrin recovery from the tissue (dry weight basis) was
12,000 to 260,000 times the initial theoretical water concentrations
(Kawatski and Schmulbach, 1971). In a model ecosystem study, residue
accunulation factors (wet weight basis) for dieldrin were determined
to be 114,935 times water concentration for the snail, 7,480 times
water concentration for algae, 6,145 times water concentration for
fish, 2,145 times water concentration for Daphnia, 1,280 times water
concentration for Elodea, 247 times water concentration for the crab,
and 1,015 times water concentration for the clam (Sanborn and Yu, 1973).
In continuous flow exposure to less than 0.1 ug/1 aldrin in water
for a three-day period, residue accunulation factors (dry weight
basis) were determined for cladocera, Daphnia magna, to be
233
-------
141,000 times water concentration; ephemeroptera, Hexagenia billineata,
31,400 times water concentration; and diptera, Chironomus sp., 22,800 times
water concentration (Johnson, et al_., 1971); for the alga, Scenedesmus
obliquus, 1,282 times water concentration after 1.5 days; 13,954 times
water concentration for Daphnia magna after 3 to 4 days, and am estimated
49,307 times water concentration for the guppy, Poecilia reticulata, after
18 days'exposure (Reinert, 1972).
In relating accumulation factors to the acceptable level of aldrin and
dieTdrin allowable in water, it is necessary to know the significance of
tissue residue levels. Data on the toxicity of ingested levels of aldrin and
dieldrin in aquatic organisms are few. In rainbow trout, Salmo gairdneri,
fed dietary dieldrin dosages of 0.36, 1.08, 3.6, and 10.8 ug dieldrin per
gram of food (ppm), brain concentrations of three amino acids associated
with ammonia detoxifying mechanisms - glutamate, aspartate, and alanine -
were significantly altered. In the two highest dosages the brain ammonia
concentration increased (Mehrle and Bloomfield, 1974). The implication is
that brain ammonia detoxifying mechanisms play an important role in main-
taining ammonia values within physiological limits, and that fish carrying
body burdens of dieldrin would be less tolerant to increased concentrations
of ammonia in water.
Some data available on terrestrial vertebrates indicate that aldrin-
dieldrin dietary levels as low as 1 ppm may produce observable effects.
In long-term feeding studies 1 ppm dieldrin affected reproduction in the
-------
Hungarian partridge (Neill, e_t aL, 1969). Slight eggshell thinning was
noted in mallard ducks fed 3 ppm dieldrin (Lehner and Egbert, 1969). Deer
were affected ty long-term feeding at 2 ppm dieldrin (Murphy and Korschgen,
1970). A guideline of 0.3 ppm has been set by the Food and Drug Adminis-
tration as an upper limit on food for human consumption.
Considering the 100,000-fold bioaccumulation of dieldrin in fish
tissue from Lake Michigan cited above (Reinert, 1970) and the FDA tissue
residue administrative guideline of 0.3 ppm as a maximum level for human
consumption, the resulting maximum water level is 0.003 ug/1. This level
is lower than the lowest measured 96-hour 1050 (7.9 ug/1 for bluegill) by
a factor of 0.0004, and lower than the 20 day LC$Q for the stonefly,
Acroneuria pacifica, by a factor of 0.02. It is therefore believed to provide
an adequate level of protection for freshwater aquatic life.
In bioassays performed on the mullet, Mugilidae sp., death occurred
for 15 percent of the test organisms at 0.5 ug/1 dieldrin. When exposed to
dieldrin at 1.35 ug/1 for four days, fish were found to have degenerative
changes in the gills and visceral tissue (Parrish, et_ aK, 1973). A
sensitive marine crab, Leptodius floridanus, exhibited delay in development
at concentrations of 1 and 0.5 ug/1 dieldrin (Epifanio, 1971). At 1 ug/1
dieldrin there was a 70 percent mortality of pink shrimp, Penaeus duorarum,
within 24 hours and the 96-hot/r LC$Q was 0.7 ug/1 for the same organism
(Parrish, et al_., 1973).
A residue accumulation factor of 800 times water concentration (wet
weight basis) was found for the estuarine mollusc, Rangia cuneata, for a
-------
36-hour exposure with a maximum accumulation factor of 2,000 times water
concentration in a 72-hour exposure to dieldrin. The clam showed "no
evidence for a cessation of, the uptake mechanism over a period of time up
to 72 hours" (Petrooelli/ et al., 1973). At a water concentration of 0.5
ug/1 it was determined that the rate of dieldrin uptake from water by crab
larvae, Leptodius floridanus, was 0.191 ppm per day from water (Epifanio,
1973). Bioconcentration factors in estuarine organisms exposed for
96-hours to dieldrin ranged from 2,400 to 21,500 for oysters; 280 to
420 for pink shrimp; 470 to 750 for grass shrimp; and 3,500 to 7,300
for sheepshead minnows (Parrish, et al., 1973). Concentration factors
in spot exposed to dieldrin for 35 days were as great as 113,000 in
liver, 11,000 in muscle and 6,000 in whole fish. Spot lost all
detectable dieldrin after 13 days in dieldrin-free sea water (Parrish,
et al., 1973). Marine phytcplankton exposed for two hours have been
shown to accumulate dieldrin (Rice and Sikka, 1973). In a study designed
to investigate the accumulation and metabolism of dieldrin by species
representing different taxonomic divisions of marine phytoplankton,
concentration factors observed were: Skeletonema costatum (Bacillariophyta),
15,882; Cyclotella nana (Bacillariophyta), 4,810; Tetraselmis chuii
(Euglenophyta), 8,588; Isochrysis galbana (Chrysophyta), 8,238; Olisthodiscus
luteus (Xanthophyta), 4,900; and Amphidinium carteri (Pyrrophyta), 982.
NO data on bioaccumulation from the environment are available for
marine fish. In the absence of such data, there is no reason to assume
that the accumulation factor for marine fish would be less than that
observed for Lake Michigan fish. Therefore, the freshwater criterion
of 0.003 ug/1 is also recommended for marine aquatic life.
-------
REFERENCES CITED:
Cairns, J., e_t al_., 1967. Effects of sublethal concentrations of
dieldrin on laboratory populations of guppies, Poecilia reticulata
Peters. Proc. Acad. Nat. Sci., Philadelphia, Pennsylvania, 119:75.
Chadwick, G.6. and R.W.' Brockson, 1969. Accumulations of dieldrin by
by fish and selected fish-food organisms. Jour. Wildlife Management,
33:693.
Epifanio, C.E., 1971. Effects of dieldrin in seawater on the development
of two species of crab larvae, Leptodius floridanus and Panopeus
herbsti. Marine Biology, 11:366.
Epifanio, C.E., 1973. Dieldrin uptake by larvae of the crab Leptodius
floridanus. Marine Biology, 19:320.
Gakstatter, J.H., 1968. Rates of accumulation of 14C - dieldrin residues
in tissues of goldfish exposed to a single sublethal dose of 14C - aldrin.
Jour. Fish. Res. Bd. Canada, 25:1797.
Henderson, C., ejt aK, 1959. Relative toxicity of ten chlorinated
hydrocarbon insecticides to four species of fish. Trans. Amer. Fish.
Soc., 88:23.
Jensen, L.D. and A.R. Gaufin, 1966. Acute and long-term effects of
organic insecticides on two species of stonefly naiads. Jour. Water
Poll. Control Fed., 38:1273.
23?
-------
Johnson, B.T., ejt al_. 1971 Biological magnification and degradation
of DDT and aldrin by freshwater invertebrates. Jour. Fish. Res.
Bd. Can. 28:705.
Kawatski, J.J. and J.C. Schmulbach, 1971. Accumulation of insecticide
in freshwater ostracods exposed continuously to sublethal concen-
trations of aldrin or dieldrin. Trans. Amer. Fish. Soc., 100:565.
Keller, A.M., e_t al_., 1974. Respondents brief, proposed findings and
conclusions on suspension of aldrin-dieldrin. FIFRA Docket No. 145.
Kahn, M.A.Q., et al_., 1972. Ir^ vivo and iji vitro epoxidation of aldrin
by aquatic food chain organisms. Bull. Environ. Contam. and Toxicol., 8:219.
Lane, C.E. and R.J. Livingston, 1970. Some acute and chronic effects
of dieldrin on the sailfin molly, Poecilia latipinna. Trans. Amer.
Fish. Soc., 3:489.
Lehner, P.N. and A. Egbert, 1969. Dieldrin and eggshell thickness in
ducks. Nature, 224:1218.
Mehrle, P.M. and R.A. Blcornfield, 1974. Ammonia detoxifying mechanisms
of rainbow trout altered by dietary dieldrin. Toxicology and Applied
Pharmacology, 27:335.
Murphy, D.A. and L.J. Korschgen, 1970. Reproduction, growth, and tissue
residues of deer fed dieldrin. Jour. Wildlife Management, 34:887.
Neill, D.D., et_al_., 1969. 58th Annual Meeting of the Poultry Science
Association. 76.
Parrish, P.R., e£ al_., 1973. Dieldrin effects on several estuarine
organisms. Proc. Assoc. S.E. Game Fish. Comm., pp. 427-434, December, 1972.
Z3S
-------
Petrocelli, S.R., et_ al_., 1973. Uptake and accumulation of an organo-
chlorine insecticide (dieldrin) by an estuarine mollusk Rangia cuneata.
Bull. Environ. Contam. and Toxicol., 10:315.
Reinert, R.E., 1970. Pesticide concentrations in Great Lakes fish.
Pesticides Monitoring Jour., 3:233.
Reinert, R.E., 1972. Accumulation of dieldrin in an alga (Scenedesmus
obliquus). Daphm'a magna. and the guppy (Poecilia reticulata). Jour.
Fish. Res. Bd. Canada, 29:1413.
Rice, C.P. and H.C. Sikka, 1973. Fate of dieldrin in selected species
of marine algae. Bull. Environ. Contam. and Toxicol., 9:116.
Sanders, H.O. and O.B. Cope. 1966. Toxicities of several pesticides
to two species of Cladocerans. Trans. Amer. Fish Soc. 95:165.
JSanborn, J.R. and C. Yu, 1973. The fate of dieldrin in a model ecosystem.
Bull. Environ. Contam. and Toxicol., 10:340.
Tarzwell, C.M. and C. Henderson, 1956. Toxicity of dieldrin to fish.
Trans. Amer. Fish. Soc., 86:245.
Walker, A.I.T., et. aK, 1972. The toxicity of dieldrin (HEOD). I.
Long-term toxicity studies in mice. Food Cosmeti Toxicol., 11:415.
7.39
-------
CHLORDftNE
CRITERIA;
0.01 ug/1 for freshwater aquatic life.
0.004 ug/1 for marine aquatic life.
The persistence1, bioaccutnulation potential
and carcinogenicity of chlordane cautions
human exposure to a minimum.
RATIONALE;
Since chlordane is a highly persistent chemical which bioaccumulates
in aquatic organisms used for human food and also is considered a
potential human carcinogen (Train, 1974), levels of chlordane in water-
ways should be kept as low as feasible. The December 24, 1975 action by
the Environmental Protection Agency in suspending the production and use
of chlordane should result in a gradual decrease in concentrations in
the environment. Such additions should not be permitted
without substantial documentation that alternatives are either infeasible
or potentially more hazardous. The persistence, bioaccumulative properties
and carcinogenic potential of chlordane should be taken into account when
determining the uses of water with measurable amounts of dhlordane.
Literature references indicate the existence of an extremely wide
range for the acute toxicity of chlordane to various species of fresh-
water fishes. As recorded, these values for 24- to 96-hour exposures
range from 5 to 3,000 ug/1 (Cardwell, et al., 1975; Katz, 1961;
Lawrence, 1950; National Technical Advisory Committee, 1968; Clemens
and Sneed, 1959; Macek, et al., 1969). Physical conditions varied
among the tests reported and, in addition, most were conducted under
2HO
-------
static conditions and actual water concentrations were not measured.
Although individual acute toxicity values may be questionable, some
tests using several species under uniform conditions indicate that
variability of experimental techniques alone cannot account for the
wide range of values published. A recent study by Cardwell, et al.,
(1975) utilizing flow-through systems and measured concentrations of
chlordane yielded 96-hour IC50 values of 37, 47, and 59 ug/1 for
fathead minnows, brook trout, and bluegills, respectively. These
values fall within the range of the majority of acute toxicity values
reported in the literature. The most sensitive fish reported was the
pike, which suffered distress after 24 hours at 5 ug/1 and 100 percent
mortality within this period at 50 ug/1 (Ludemann and Neumann, 1962).
Published acute toxicity values for U.S. freshwater invertebrates
are similar to those for fishes and range from 4 to 10,000 ug/1
(Cardwell, et al., 1975; Naqyi and Ferguson, 1969; National Technical
Advisory Committee, 1968; Sanders 1969, 1972). Daphnia magna and
Hyallela azteca tested under flow-through conditions yielded 96-hour
LC50 values of 28 and 97 ug/1 (Cardwell, et al., 1975), falling within
the range of the majority of values reported for acute toxicity to
invertebrates. The most sensitive species reported was the glass
shrimp, Palaemonetes kadiakensis, with 96-hour LC50 values of 10
and 4 ug/1, respectively, in static versus flow-through bioassays
(Sanders, 1972). Ludemann and Neumann (1962) found that Chironomus
larvae had a 24-hour static test LC50 value of 10 ug/1.
-------
Cardwell, et al. (1975), conducted long-term flow-through studies
which included reproduction and survival of progeny with several aquatic
species. Statistically significant detrimental effects could not be
measured at concentrations of 0.5 ug/1 (bluegills), 0.7 ug/1
(Chironomus No. 51, a midge), 5 ug/1 (Hyallela azteca) and 12 ug/1
(Daphnia magna). Survival of brook trout larvae to 12 days was reduced
by 35 percent at 0.3 ug/1 (the lowest concentration tested) but overall
effects were decreasing at this level.
Available acute toxicity values indicate that individual species
of both fishes and invertebrates vary greatly in sensitivity to chlordane
but both groups are within the same general range. Most of these acute
toxicity values are also nearer to the lower limits of this sensitivity
range. Data for the application factors derived using acute and sub-
chronic toxicity concentrations are limited to the study by Cardwell,
et al. (1975). These factors were 0.009 and less than 0.008 for the
fishes and 0.381 and 0.055 for the invertebrates.
Nominal pesticide concentrations in the water were reported for
these tests. The true concentrations of chlordane may have been somewhat
lower as Cardwell, et al. (1975), found that concentrations could not be
maintained above half of nominal even when the test water was continually
replaced. Therefore, the water quality criterion should be set lower
than the calculated value using the experimentally determined application
factpr ( 0.009.X 5.0 ug/1 = 0.0^5 ug/l). For this reason the calculated
value for chlordane is xsssui&Bskx&JBKKx lowered to a recommended concentration
1n freshwater that should not exceed 0.01 ug/1. If the
proposed freshwater water quality criterion of 0.01 is not
exceeded it is estimated that levels reached in freshwater
-------
fishes should seldom be greater than 1.0 rag/kg- Based on present
knowledge/ it is not anticipated that a concentration of this magnitude
would be detrimental to piscivorous birds.
Michael, et al. (1956), found that the time required to kill one-
half of the brine shrimp, Artemia salina, nauplii exposed to 10 ug/1
was 2-3 hours. Butler, et al. (1960), determined that 24 hours' exposure
of the oyster, Grassestrea virginica, to 10 ug/1 chlordane produced growth
inhibition.
Kbrn and Earnest(1974) found the 96-hour LC50 of chlordane to be
11.8 ug/1 for the striped bass, Morone saxatilis. Butler (1963) reported
the following 48-hour LC50 values: brown shrimp, Penaeus aztecus —
4.4 ug/1; juvenile blue crabs, Callinectes sapidus — 480 ug/1; and
juvenile white mullet, Mugil curema — 5.5 ug/1. Parrish, et al. (In Press),
reported the following 96-hour I£50 values: pink shrimp, Penaeus duorarum —
0.4 ug/1; grass shrimp, Palaemonetes pugio — 4.8 ug/1; sheepshead minnow,
Cyprinodon variegatus — 24.5 ug/1; and the pinfish, Lagodon rhomboides —
6.4 ug/1. The BC50 for shell deposition by the eastern oyster, Crassostrea
virginica, was found to be 6.2 ug/1. The most sensitive marine species
tested was the pink shrimp, Penaeus duorarum, for which a 96-hour LC50
of 0.4 ug/1 was reported.
It appears that reported concentrations of "total chlordane" in
fishes have been calculated from measurements of one or two of its
major components under the assumption that the ratio of components was
similar in both the parent compound and the tissue residues. More
-------
sophisticated analytical techniques currently available indicate that
this assumption is in error and, in addition/ the components in tissue
residues may have been incorrectly quantified. Fishes can concentrate
chlordane directly from .water by a factor of 1000 to 3000 times and
invertebrates may concentrate to twice this magnitude (Cardwell,, et al.,
1975). Data on the bioaccumulation of chlordane by estuarine organisms
have been reported by Parrish, et al. (In Press). In the 96-hour test,
chlordane concentration factors were 3,200 to 8,300 in oysters, Grassestrea
virginica ; 4,000 to 6,000 in pink shrimp, Penaeus duprarum ; If900 to
2,300 in grass shrimp, Palaemonetes pugio ; 12,600 to 18,700 in pinfish,
Lagodon rhomboides . Schiimel, et alv (In Press), exposed estuarine
fishes to trans-chlordane for 96 hours and reported concentration factors
ranging from 3,700 to 6,300 in edible portions. Reported concentrations
of 10 to 100 ugAg were common in fish samples obtained throughout the
U.S. and a few residues above 1000 ug/kg were observed (Henderson, et al.,
1969 and 1971). Corresponding water concentrations are not available, but
measurements above nanogram/liter levels have seldom been found in natural
waters except near the discharges of manufacturing and formulating
operations (Barthel, et al., 1969; Casper, 1967; Godsil and Johnson, 1968).
Therefore, allowing for errors in residue measurements and lack of direct
correlation, the possibility exists that fishes can concentrate chlordane
up to 100,000 times the ambient water concentration. High rates of
accumulation could occur in higher trophic level organisms through multiple
steps in the food chain although evidence directly related to chlordane
is not available to support this hypothesis.
-------
Based on the persistence, bioaccumulation potential and
carcinogenicity of chlordane, an application factor of .01 is applied
to the most sensitive marine aquatic species, the pink shrimp. This
results in a marine criterion of .004 ug/1.
-------
REFERENCES CITED:
Barthe1, W.F., et al., 1969. Pesticide residues in sediments
of the lower Mississippi River and its tributaries.
Pest. Mon. Jour. 3:8.
Butler, P.A., et al., I960. Effect of pesticides on oysters.
Proceedings of the National Shell Fisheries Association.
51:23.
Butler, P.A., 1963. Commercial Fisheries Investigations.
In: Pesticide Wildlife Studies. U. S. Dept. of the
Interior. Washington, D. C. Circular 167.
Cardwell, R. D., et al., 1975. Acute and chronic toxicity of
chlordane to fish and invertebrates. EPA Ecological
Research Series. In preparation.
Casper, V. L., 1967. Galveston Bay pesticide study - water
and oyster samples analyzed for pesticide residues
following mosquito control program. Pest. Mon. Jour.
1:13.
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Clemens, H. P. and K. E. Sneed, 1959. Lethal doses of
several commercial chemicals for fingerling channel
catfish. U. S. Pish. Wildl. Serv. Spec. Set. Rep. Fish,
316.
Godsil, R. S. and W. C. Johnson, 1968. Residues in fish,
wildlife, and estuaries-- pesticide monitoring of the
aquatic biota at the Tule Lake National Wildlife Refuge,
Pest. Mont. Jour. 1:21.
Henderson, D. , et al. , 1969. Organochlorine insecticide
residues in fish. Pest. Mon. Jour. 3:145.
Henderson, C., et al. , 1971. Organochlorine insecticide
residues in fish - fall, 1969, National Pesticide
Monitoring Program. Pest. Mont. Jour. 5:1.
Katz, M. , 1961. Acute toxicity of some organic insecticides
to three species of salmon ids and to the threespine
stickleback. Trans. Amer. Fish. Soc. 90:264.
Koro, S. and R. Earnest, 1974. Acute toxicity of twenty
insecticides to striped bass, Morone saxatilis . Calif.
Fish and Game, 60:128.
-------
Lawrence, J. M., 1950. Toxicity of some new insecticides to
several species of pondfish. Prog. Fish. Cult. 12tl41.
Ludemann, D., and H. Neumann, 1962. Uber die wirkung der
neuzeitlichen kontakinsektizide auf die tiere des
susswassers. Anzeiger fur Schadlingskunde und
Pflanzenschutz, 35:5.
Macek, K. B., et al., 1969. Effects of temperature on the
susceptibility of bluegills and rainbow trout to
selected pesticides. Bull. Environ. Contam. Toxicol.,
4:174.
Michael, A.S., et al., 1956. Arteraia aalina as a test
-» -_•> ___->_^^«^ _______
organism for bioassay (of insecticides). Science
123:464.
Uaqvi, S. M. and D. E. Ferguson. 1969. Pesticide tolerances
of selected freshwater invertebrates. Jour. Miss. Acad.
Sci. 14:121.
-------
National Technical Advisory Committee, 1968. Wa£er Quality
Criteria. Fed. Water Poll. Control Admin. U.S. G.P.O.
Washington, D.C.
Parrish, P. R., et al., (In Press), Chlordane: Effects on several
estuarine organisms. Jour. Toxiool. Environ. Health.
Sanders, H. 0.-, 1969. Toxicity of pesticides to the
crustacean, Garomarus lacustris. Bur. Sport Pish, and
Wildl. Tech. Paper No. 25.
Sanders, H. O., 1972. Toxicity of some pesticides to four
species of malocostracan Crustacea. Bur. Sport Fish.
and Wildl. Tech. Paper No. 66.
Schimrnel, S. C., et alv (In Press). Heptachlor: Toxicity to and
uptake by several estuarine animals. Jour. Toxicol. Environ.
Health.
Train, R. E., 1974. Pesticide products containing heptachlor
or chlordane: Intent to cancel registrations, Federal
Register, Vol. 39, No. 229, Tues, Nov. 26, 1974.
P.41299.
-------
CRITERIA!
Chlorophenoxy Herbicides
2, 4-D; 2, 4, 5-TP
2, 4-D
2,4,5-TP
100 ug/1
10 ug/1
for domestic water supply (health)
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 miTd injury fn
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
coccidioidomycosis, the patient had no side effects from 18
intravenous doses during 33 days; each of the last 12 doses in
the series was 800 mg (about 15 mg/kg) or more, the last being
2000 mg (about 37 mg/kg) (Seabury, 1963). A nineteeth and final
dose of 3600 mg (67 mg/kg) produced mild symptoms.
Table $ illustrates the derivation of the criteria for the
two chlorophenoxy herbicides most widely used. The long-term
no-effects levels (rag/kg/day) are listed for the rat and the dog.
These 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.
-------
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-------
REFERENCES CITED:
Kraus, as cited by Mitchess, J.W., R.E. Hogson, and C.R.
Gaetjens, 1946. Tolerance of farm animals to feed
containing 2, 4-dichlorophenoxyacetic acid. Jour;
Animal. Sci., 5s 226.
Lehman, A.J., 1965. Summaries of pesticide toxicity.
Association of Food and Drug Officials of the U.S.,
Topeka, Kansas, pp. 1-40.
Palmer, J.S. and R.D. Radeleff,' 1964. The toxicologie
effects of certain fungicides and herbicides on sheep
and cattle. Ann. N.Y. Acad. Sci., lilt 729.
Seabury, J.H., 1963. Toxicity of 2, 4-dichlorophenoxyacetic
acid for man and dog. Arch. Envir. Health, 7s 202.
-------
DDT
CRITERION:
0.001 ug/1 for freshwater and marine aquatic life.
The persistence, bioaccunulation potential
and carcinogenicity of EOT cautions human
exposure to a minimum.
RATIONALE:
In general, DDT (1,1,l-trichloro-2,2-bis(p-chlorophenyl)ethane)
refers to DDT and its metabolites. Acute toxicity to mammals generally
1s low; however, aquatic organisms exhibit sensitivity to this pesticide
at levels in micrograms per liter. Such levels range from a 96-hour
LC5Q of 0.24 ug/1 for the crayfish, Orconectes nais (Sanders, 1972),
to a 96-hour LC^Q of 2 ug/1 for the largemouth bass, Micropterus salmoides
(Macek and McAllister, 1970), to a 96-hour LC5Q of 27 ug/1 for the
goldfish, Carassius auratus (Henderson, ejt aj_., 1959).
Since DDT is a highly persistent chemical which bioaccumulates in
aquatic organisms used for human food and also is considered a potential
human carcinogen (Train, 1975), levels of DDT in waterways should be kept as
low as feasible. Action by the Environmental Protection Agency in
suspending the production and use of DDT should result in a gradual
decrease in concentrations in the environment, such
additions should not be permitted without substantial documentation that
alternatives are either infeasible or potentially more hazardous. The
persistence, bioaccumulative properties and carcinogenic potential of
-------
DDT should be taken Into account when determining the uses of water
containing measurable amounts of DDT.
to 4-day LCBO's for freshwater aquatic invertebrates and fish
exposed to DDT in water generally have ranged in the low microgram per
liter levels with the invertebrates being somewhat more sensitive
(Sanders, 1969; Sanders, 1972; Sanders and Cope, 1966 and 1968
Henderson, et. aj_, 1959; Macek and McAllister, 1970). The most sensitive
freshwater organism for which there are data is the crayfish, Qrconectes
nais. which had a 96-hour LC50 of 0.24 ug/1 (Sanders, 1972). Of marine
organisms the most sensitive are the shrimp, Penaeus duorarum and P.
setiferus. for which no survival was observed at 0.12 ug/1 after 28 days1
exposure and 30 percent mortality observed at 0.05 ug/1 after 56 days
(Nimmo, et al_. , 1970)
DDT will accumulate in the food chain. Field data where pesticide
levels in both the water and aquatic organisms have been measured depict
the quantity attributable to this bioaccumulation. A residue accumulation
of up to two million times was calculated for fish from the field data
of Reinert (1970). The measured DDT water concentration in his study of
Lake Michigan was in the nanogram per liter range. This accumulation
is 20 times greater than that observed in the National Water Quality
Laboratory, Duluth, Minnesota, where residue accumulation of 100,000 times
•
was observed. If one uses a two million times residue accumulation factor
and a DDT water concentration of 0.002 ug/1 an expected DDT body burden
of 4 mg/kg (ppm) in fish would result.
Other studies show that measurable quantities of DDT can be found
in natural waters of North America. DDT averaged 0.02 ug/1 in the San
Diego River, California, in the spring-fall dry weather flow of 1973
(Young and Beeson, 197^ DDT in ocean water along the west coast has
ranged from 0.0023 to 0.0056 ug/1 (Cox, 1971). Rivers flowing
-------
Into estuaries (Brazos and Colorado in Texas) each have a two-year
average of ,0.03 ug/1 (Manlgold and Schulze, 1969).,
I it-laboratory studies, Hansen and Wilson (1970) found that the bio-
accumulation factor from water to pinfish, Lagodon rhomboides, was
10,000 times and to Atlantic croaker, Micropogon undulatus, 38,000 times,
or an average of about 25,000 times for these fishes. The discrepancy
^between laboratory and field data jr*# fcC. due to the,many additional
tiroghic levels involved in field exposures.
In controlled studies, Heath, et^aj[. (1969) determined that DDE, in
concentrations of 10 and 40 ppm DDE in dry feed, impaired reproductive
success of penned mallards, Anas platyrhynchos. Eggshells of birds
exposed to these concentrations were 13 percent thinner, with 25 percent
of the eggs showing cracking after one month. DDT induced thinning of
shells at a concentration of 25 ppm with an 18 percent cracking of the
shells. Since DDT metabolizes to DDE, eggshell thinning may be partly
due to DDE. The association of DDE residues with eggshell thinning
was shown for the brown pelican, Pelecanus occidental is (Blus, et al.,
1972). The level of DDE in the eggs which did not produce a significant
effect, I.e., thinning, was estimated to be 0.5 mg/kg. However, from the
data presented, 2.0 mg/kg represents a conservative estimate of the no-
effect level in the eggs.
Feeding studies .have shown that black ducks fed DDT in food produced
eggs containing residues of about ten-fold the DDT in the diet. Further-
more, a continuous diet of 3.0 mg/kg (wet weight) in natural food
-------
adversely affected reproduction (Longcore, et a^., 1971). Extrapolating
the egg concentration data in ducks to pelicans^an estimate can be made
that the diet of pelicans should contain no more than 0.1 the estimated
(2.0 mg/kg) no-effect level of DDT in eggs, or about 0.2 mg/kg.
Based on the considerations of bioaccumulation potential in fish,
likelihood of conversion to DDE, and dosage levels known to adversely
affect birds, 1t 1s recommended that DDT concentrations 1n water should
not exceed 0.001 ug/1.
-------
REFERENCES CITED:
Blus, L.J., et al_., 1972. Further analysis of the logarithmic relation-
ship of DDE residues to nest success. Nature, 240:164.
uCox, J.L., 1971. DDT residues in sea water and particulate matter in
the California current system. Fishery Bulletin, 69:443.
Hansen, D.J. and A.J. Wilson, Jr., 1970. Residues in fish, wildlife
and estuaries. Pesticides Monit. Jour., 4:51.
Heath, R.G., §t aK, 1969. Marked DDE impairment of mallard reproduction
in controlled studies. Nature, 224:47.
Henderson, C., et al., 1959. Toxicity of organic phosphorus and
chlorinated hydrocarbon insecticides to fish. Biol. Probl. Water Poll.
Trans., Seminar, Robert A. Taft San. Eng. Center, Tech. Report W60-3:76.
Longcore, J.R., et a]_., 1971. DDE thins eggshells and lowers'reproductive
success of captive black ducks. Bull. Environ. Contam. Toxicol., 6:485.
Macek, K.J. and W.A. McAllister, 1970. Insecticide susceptibility of
some common fish family representatives. Trans. Amer. Fish. Soc., 99:20.
Manigold, D.E. and J.A. Schulze, 1969. Pesticides in water. Pesticides
Monit. Jour., 3:124.
Nimmo, D.R., et al_., 1970. Localization of DDT in the body organs of
pink and white shrimp. Bull. Environ. Contam. Toxicol., 5:333.
35?
-------
Reinert, R.E., 1970. Pesticide concentrations In Great Lakes fish.
Pesticides Monlt. Jour., 3:233.
Sanders, H.O., 1969. Toxiclty of pesticides to the crustacean Gammarus
lacustris; U.S. Bureau of Sport Fisheries and Wildlife, Tech. Paper No. 25.
Sanders, H.O., 1972. The toxicities of some insecticides to four species
of malacostracan Crustacea. U.S. Bureau of Sport Fisheries and Wildlife,
Tech. Paper No. 66.
Sanders, H.O. and O.B. Cope, 1966. The toxicities of several pesticides
to two species of cladocerans. Trans. Amer. Fish. Soc., 85:165.
Sanders, H.O. and O.B. Cope, 1968. The relative toxicities of several
pesticides to naiads of three species of stoneflies. Limnol. Oceanog.,
13:112.
Train, R., 1975. Order and determination of the administration that
reconsideration of the Agency's prior order of cancellation of DDT
for use on cotton is not warranted, Federal Register, 40:15934.
Young, D.R. and T.C. Beeson, 1974. Inputs and distributions of chlorinated
hydrocarbons 1n three Southern California harbors. Southern California
Coastal Water Res. Project, Tech. Paper TM-214.
25?
-------
DEMETON
CRITERION!
0.1 ug/1 for freshwater and marine aquatic life
RATIONALE;
Static LC bioassays yielded toxicity values for the organo-
phosphorus pesticide, demeton, fpr carp, goldfish, fathead minnow,
channel catfish, guppy, rainbow trout, and bluegill, ranging from 70 ug/1
to 15, 000 ug/1 (Henderson and Pickering, 1958; Ludemann and Neumann,
1962; Macek and McAllister, 1970; McCann and Jasper, 1972; Pickering,
et^aJL., 1962). Results of these tests demonstrate an apparent sharp
division in species sensitivity, with bluegill, Lepomis macrochirus; rain-
bow trout, Salmo gairdneri; and guppy, Poecilia reticulata, being
susceptible to lower concentrations while the remaining species were
comparatively resistant. In the 96-hour exposures toxicity did not in-
crease 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 LC of 70 ug/1 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 and 1961) the maximum inhibition of brain acetylcholin-
esterase (AChE) was not reached. The lowest levels of /\chE occurred
after 24 to 48 hours. It was demonstrated that maximum inhibition could
last as long as two weeks after exposure, and subsequent recovery to
-------
levels approaching normal took many more weeks. Weiss (1958)
reported a significant increase in mortality of fathead minnows exposed
for a second time to the organophosphate, Sarin, before the fish had
recovered normal brain -AChE levels. The resistance of fully recovered
fish was equal to that of previously unexposed controls. Weiss and
Gakstatter (1964a) reported no significant inhibition of brain AChE in
bluegills, goldfish, and shiners, Notemigonus crysoleucas, following
15-day exposures to demeton at continuously replenished, nominal concen-
centrations of Lug/1.
Acute toxicity values reported for invertebrates range from 10 to
100,000 ug/1 (Ludemann and Neumann, 1962; Sanders, 1972). In general,
molluscs and tubifex worms were very resistant while the smaller
crustaceans and insect larvae were susceptible. Ludemann and Neumann
(1962) reported that Chironomus plumosus larvae were the most sensitive
species they tested. A 24-hour exposure at 10 ug/1 produced undefined
effects while 100 percent were killed at 1000 ug/1. Calculated LCgQ data
for invertebrates apparently are limited to a single, nominal concentration
static exposure of Gammarus fasciatus (Sanders, 1972). These 24- and
96-hour LCgQ values are reported as 500 and 27 ug/1, 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 sub-
acute concentrations of organophosphates.
-------
There are few data oft the toxicity of demeton to marine organisms.
Butler (1964) reported a 48-hour BC50 of 63 ug/1 for the pink shrimp,
Penaeus duorarum, and a 24-hour 1C of 550 ug/1 for the spot, Leiostomus
xantjgurus.
Chronic demeton toxicity data for freshwater organisms are not currently
available. Since there are no data 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 ten
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/1
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.
-------
REFERENCES CITED;
Butler, P.A., 1964. Commercial fishery investigations. In: Pesticide
Wildlife Studies, 1963. p. 5-28. U.S. Fish and Wildlife Service
Circular 149. Washington, D.C.
, Henderson, C. and Q. H. Pickering, 1958. Toxicity of organic
phosphorus insecticides to fish. Trans. Amer. Fish. Soc., 87:39.
Ludemann, D. and H. Neumann, 1962. Uber die wirkung der neuzeitlichen
kontakinseckitizide auf die tiere des susswassers. Anzeiger fur
Schadlingskunde und Pflanzenschutz, 35:5.
Macek, K. J. and W. A. McAllister, 1970. Insecticide susceptibility of
some common fish family representatives. Trans. Amerl Fish. Soc.,
99:20.
McCann, J. A. and R. L. Jasper, 1972. Vertebral damage to bluegills
exposed to acutely toxic levels of pesticides. Trans. Amer. Fish. Soc.,
101:317.
Pickering, Q.H., et^ aL , 1962. The toxicity of organic phosphorus
insecticides to different species of warmwater fishes. Trans. Amer.
Fish. Soc., 91:175.
Sanders, H. O., 1972. Toxicity of some insecticides to four species of
malacostracan crustaceans. U. S. Dept. of the Interior, Washington,
D. C., Bureau of Sport Fisheries and Wildlife, Tech. Paper No. 66.
pp. 3-19.
Weiss, C.M., 1958. The determination of cholinesterase in the brain
tissue of three species of freshwater fish and its inactiviation in vivo.
Ecology, 39:194.
-------
/Weiss, C.M., 1959. Response of fish to sub-lethal exposures of
organic phosphorus insecticides. Sew. and Ind. Wastes, 31:580.
v Weiss, C.M., 1961. Physiological effect of organic phosphorous
insecticides on several species of fish. Trans. Amer. Fish. Soc.,
90:143.
" Weiss, C.M. andJ.H. Gakstatter, 1964a. Detection of pesticides in
water by biochemical assay. Jour. Water Poll. Cont. Fed., 36:240.
Weiss, C.M. and J.H. Gakstatter, 1964b. The decay of anticholinesterase
activity of organic phosphorus insecticides on storage in waters of
different pH. Advances in Water Poll. Research, 1:83-95.
-------
ENDOSULFAN
CRITERIA:
0.003 ug/1 for freshwater aquatic life;
0.001 ug/1 for marine aquatic life.
RATIONALE:
The acute toxicity of endosulfan (also known as thiodan) to
different fish species varies widely. Macek, e_t al_. (1969) exposed
rainbow trout, Sal mo gairdnerl, to endosulfan at three temperatures
and computed 24-hour and 96-hour LC5Qs. At 1.6° C, 7.2° C, and
12.7° C the 24-hour LCSO's were 13, 6.1, and 3.2 ug/1, respectively.
The corresponding 96-hour U^g values were 2.6, 1.7, and 1.5 ug/1.
Schoettger (1970), however, reports the 96-hour LC5Q for rainbow trout
to be 0. 8 ug/1 at 1. 5° C and 0. 3 ug/1 at 10° C. He also determined the
96-hour LC for the western white sucker, Castostomus commersoni, to
50 o o
be 3. 5 ug/1 at 10 C and 3.0 ug/1 at 19 C.
A massive fish kill in the Rhine River was attributed to a maximum
concentrations of 0. 7 ug/1 endosulfan (Greve and Wit, 1971). The 96-hour
o
LC at 20 C in a static bioassay using the guppy, Poecilia reticulata,
50
was 4. 2 ug/1 based on the computed concentration. The measured
concentration was only 0.2 ug/1 (Herzel and Ludemann, 1971).
The 24-, 48-, and 96-hour LC for the amphipod, Gammarus
50s
lacustris, were found to be 9.2, 6.4, and 5.8 ug/1 (Sanders, 1969).
Sanders and Cope (1968) determined the 24-, 48-, and 96-hour LC50's
for naiads of the stonefly, PteronaVcys
-------
call formca, to be 24, 5.6, and 2.3 ug/1, respectively. The 96-hour
LC5Q for Gammarus fasciatus was found to be 6.0 ug/1 (Sanders., 1972).
No data are available on the levels to which endosulfan could be
expected to accumulate in tissues of aquatic organisms at various water
concentrations. Residues in fish are not anticipated to pose a hazard
to fish-eating predators because of endosulfan's low oral toxicity to
birds (Heath, e£ al_., 1972) and mammals (Lindquist and Dahm, "1957). The
U.S. Food and Drug Administration has not set allowable limits for
endosulfan in edible fish tissues.
A 0.01 application factor applied to the lowest measured 96-hour
LCsg for the rainbow trout (which appears to be the most sensitive
native freshwater organism) results in a freshwater criterion of 0.003 ug/1
Portman and Wilson (1971) determined the acute toxicity of
endosulfan to a marine fish and several invertebrates by means of static
bioassays. The 48-hour 1050 for thepogge. a fish » Agonus cataphractes,
was 30 ug/1; the 48-hour LC$Q for a mussel, the European cockle,
Cardium edule, was greater than 10,000 ug/1; the 48-hour LC50 for the
shrimp, Crangon crangon, was 10 ug/1.
Butler (1963) reported a 48-hour EC50 death or loss of equilibrium
of 0.2 ug/1 for the brown shrimp, Penaeus aztecus, a 48-hour EC5Q for
juvenile blue crabs, Callinectes sapidus, of 35 ug/1; and a 48-hour
EC5Q of 0.6 ug/1 for juvenile white mullet, Mugil curema. A concen-
tration of 65 ug/1 resulted in a 50 percent decrease in shell growth
of the American oyster, Grassestrea virgim'ca, at 28° C and a salinity
Hi*
-------
of 22 o/oo. Korn and Earnest (1974) report a 96-hour LC$Q of 0.1 ug/1
for the striped bass, Morone saxatilis.
Use of an application factor of 0.01 times the 96-hour LC50 of
the most sensitive marine organism tested, the striped bass, results
in a marine criterion of 0.001 ug/1.
-------
REFERENCES CITED:
"Butler, P.A., 1963. Pesticide wildlife studies. A review of Fish and
Wildlife Service investigations during 1961 and 1962. USDI, Fish and
Wildlife Service Circ. No. 167.
frGreve. P.A. and S.L. Wit, 1971. Endosulfan in the Rhine River. Jour.
Water Poll. Cont. Fed., 43:2338.
Heath, R.G., et aL., 1972.
Cotparative dietary toxicities of pesticides to birds. U.S. Dept.
of the Interior,. Bureau of Soort Fisheries and Wildlife. Special
Scientific Report - Wildlife No. 152. Washington D.C.. 57 p.
i/Herze*!, E. and D. Ludemann, 1971. Behavior and toxicity of endosulfan
in water under various experimental conditions. Z. Angew. Zool., 58:57.
Lindquist, D. and P.A. Dahm, 1957. Some chemical and biological experi-
ments with thiodan. Jour. Econ. Entomol., 50:483.
i/Korn, S. and R. Earnest, 1974. Acute toxicity of twenty insecticides to
striped bass, Morone saxatilis. Calif. Fish and Game, 60:128.
Macek, K.J., et al., 1969. Effects of
temperature on the susceptibility of bluegills and rainbow
trout to selected p^st^ci^es. Bull. Rnviron. Contan. Ttaxicol.,
4:174.
^Portman, J.E. and K.W. Wilson, 1971. The toxicity of 140 substances to
the brown shrimp and other marine animals. Ministry of Agriculture,
Fisheries and Food, Fisheries Laboratory, Burnham-on-Couch, Essex,
England, Shellfish Intonation LaifUt No. 22, 12 pp.
-------
Sanders, H.O., 1969. Toxicity of pesticides to the crustacean,
Gammarus lacustris. U.S. Dept. of the Interior, Bur. Sport Fish.
Wildl., Tech. Paper 25.
Sanders, H.O., 1972. Toxicity of some insecticides to four species
of malacostracan crustaceans. U.S. Dept. of the Interior, Bur. of
Sport Fish, and Wildl., Washington, D.C., Tech. Paper 66.
Sanders, H.O. and O.B. Cope, 1968. The relative toxicities of several
pesticides to'naiads of three species of stoneflies. Limnol.
Oceanog., 13:112.
Schoettger, R.A., 1970. Toxicology of thiodan in several fish and
aquatic invertebrates. Bur. Sport Fish. Wildl., Investigations in
Fish Control, 35. U.S. Government Printing Office, Washington,D.C.
-------
ENDRIN
CRITERIA:
0.2 ug/1 for domestic water supply (health);
0.004 ug/1 for freshwater and marine aquatic life.
RATIONALE:
The highest level of endrin found to have minimal or no long-term
effects in the most sensitive animal tested, the dog, is 1.0 mg/kg in the
diet or 0.02 mg/kg of body weight/day (Treon, ejt al_., 1955). Where adequate
human data are not available for corroboration of the animal results, the
total "safe" intake level is assumed to be 1/500 of the "no effect"
or "minimal effect" level reported for the most sensitive animal tested.
Applying the available data and based upon the assumption*that 20
percent of the total intake of endrin is from drinking water, that the
average person weighs 70 kg and consumes 2 liters of water per day,.the
formula for calculating a criterion is .02 mg/kg x 0.2 x 70 kg x 1/500 x
1/2 =..00028 mg/1 thus deriving the criterion level for domestic water
supply of 0.2 ug/1.
Toxicity data for the flagfish, Jordanella floridae, indicate that
the "safe" concentration as determined in a long-term exposure involving
reproduction is about 0.30 of the 96-hour LCgg (Hermanutz, 1974). Ninety-
six-hour LC50's for some of the sensitive freshwater fish tested are as
follows: bluegills, 0.6 ug/1 (Henderson, et aK, 1959); rainbow trout,
0.6 ug/1 and coho salmon, 0.5 ug/1 (Katz, 1961); juvenile striped bass
(freshwater and estuarine life-phase), 0.094 ug/1 (Korn and Earnest, 1974);
-------
cutthroat trout, 0.113 ug/1 (Post and Schroeder, 1971). Therefore, the
estimated "safe" water concentrations for these sensitive fish, based
on the flagfish application factor is approximately 0.03 to 0.18 ug/1.
Stonefly naiads appear to be the most sensitive invertebrates
tested. Jensen and Gaufin (1966) found the 30-day LC50 for the naiad,
Acroneuria pacifica, to be 0.035 ug/1. Based on these data, the safe
water concentration should be less than 0.035 ug/1 to fully protect stone-
flies throughout their entire life cycle, as well as organisms more sensitive
than those that have been tested.
Fathead minnows, Pimephales promelas, exposed to 0.015 ug/1 in the
ambient water had total body residues 10,000 times greater than the water
concentrations (Mount and Putnicki, 1966). Residue accumulation up to
10,000-fold was observed in flagfish exposed for 60 days to several
concentrations between 0.05 and 0.3 ug/1 (Hermanutz, 1974). Concentration
factors for estuarine organisms exposed to endrin for 96-hours were a
maximum of 1,600 in oysters, 1,100 in pink shrimp, 860 in grass shrimp,
4,500 in sheepshead minnows and 2,500 in the sailfin molly (Schimmel,
et^ aJL, 1975 ). Johnson (1967) calculated the concentration in adult
medeka tissue to be 17,000 to 26,000 times the water concentration. It
fs quite possible that some fish would accumulate endrin to 30,000 times
water concentration. This degree of accumulation is based only on
observations of uptake directly from the water and does not allow for
accumulation via the food chain or significantly higher accumulation
rates possible in other, untested fish species.
Endrin has been found to be eliminated quickly after termination
of exposure. Channel catfish tissue residues were reduced *• ' '
-------
from 0.41 to 0.02 ug/g (about 95
percent reduction) within 13 days after the addition of endrin to the
water was stopped (Argyle, ejt a\_., 1973). Marine spot tissue residues
of 78.0 ppb were reduced below detection levels within 13 days (Lowe,
1966). Recent unpublished data showing that flagfish eliminated about
95 percent in five days support this observation. This apparent ability
to excrete endrin readily may reduce the threat of extremely high
residue accumulations.
Levels of 0*1 ug/1 have been shown to be 100 percent lethal
to the marine spot; Leiostomus xanthuras, within five days, while
approximately half the population survived at;a level of 0.075 ug/1
after 19 days (Lowe, 1966).
Davis and Hidu (1969) reported the 48-hour TLm of American oyster
eggs to be 0.79 mg/1 and the 14-day TLm for larvae to be greater than
10 mg/1.
Eisler (1969) determined
the following 96-hour LC50 values: Sand shrimp, Crangon septemspinosa -
1.7 ug/1; grass shrimp, Palaemonetes pugio - 1.8 ug/1; and hermit crab,
Pagurus longicarpus - 12 ug/1. The following 96-hour LC$Q values were
reported by Eisler (1970): Atlantic silverside, Menidia menidia -
0.05 ug/1; blue head, Thalassoma bifasciatum - 0.1 ug/1; striped killifish,
Fundulus majalis - 0.3 ug/1', striped mullet, Mugil cephalus - 0.3 ug/1;
American eel, Anguilla rostrata - 0.6 ug/1; mummichog, Fundulus
heteroclitus - 0.6 ug/1; and Northern puffer, Sphoeroides maculatus -
3.1 ug/1. Tt* W~ho«r LC50 for striped bass was O.Q94 us/1
-------
(Kbrn and Earnest, 1974),shiner perch, Cymatogastes aggregate, 0.12
-og/1 and dwarf perch, Micronetrus minimus, 0.13 ug/1 (Earnest and Benville,
1972). Ninety-six hour LCSO's based on measured concentrations were
0.63 ug/1 for sailfin mollies, Poecilia latipinna, 0.38 ug/1 for sheep-
shead minnow, Cyprinodon variegatus, 0.63 ug/1 for grass shrimp,
Palaemonetes pugio, and 0.037 ug/1 for pink shrimp, Penaeus duorarum
(Schiirmel, et al., 1975). Data on the effects of endrin in an exposure
throughout - the entire life cycle of the sheepshead minnow indicate
that a "safe" concentration to protect for reproductive effects would
be about 0.3 of the 96-hour LC50 (Hansen, In Press). An application factor of
0.01 of the 96-hour LC50 may be over protective. Therefore, for endrin
this factor should be 0.1. Based on this, the criterion for salt water
should be 0.1 of the 96-hour LC50 for pink shrimp: 0.004 ug/1. This
criterion should also be protective of freshwater aqaatic life.
-------
REFERENCES CITED:
Argyle, R.L., e_t al_., 1973. Endrin uptake and release by fingerling
channel catfish (Ictalurus punctatus). Jour. Fish. Res. Bd. Can., 30:1743.
Davis, H.C. and H. Hidu, 1969. Effects of pesticides on embryonic development
of clams and oysters and on survival and growth of the larvae. U.S.
Dept. of the Interior, Fish, and Wild!. Serv, Fishery Bull. 67:393.
Earnest, R, and P.E. Benville, Jr. 1972. Acute toxicities of
four organochlorine insecticides to two species of surf perch.
Calif. Fish Game 58(2):127-132.
Eisler, R., 1969. Acute toxicities of insecticides to marine decapod
crustaceans. Crustaceana, 16:302.
Eisler, R., 1970. Acute toxicities of organochlorine and organophosphorus
insecticides to estuarine fishes. U.S. Dept. of the Interior, Bur. Sport
Fish, and Wildl., Tech. Paper 46.
Hansen, D.J. In Press. Techniques to assess the effects of toxic
organics on marine organisms. EPA Ecol. Res. Ser.
Henderson, C., e£al_., 1959. Relative toxicity of ten chlorinated hydro-
carbon insecticides to four species of fish. Trans. Amer. Fish. Soc., 88:23.
Hermanutz, R., 1974. Northwest water quality report. National Water Quality
Laboratory, Duluth, Minnesota.
Jensen, L.D. and A.R. Gaufin, 1966. Acute and long-term effects of
organic insecticides on two species of stonefly naiads. Jour. Water
Poll. Cont. Fed., 38:1273.
-------
Johnson, H.E., 1967. The effects of endrin on the reproduction of a
freshwater fish (Oryzias latipes). Ph.D. Thesis, University of
Washington, Seattle, 135 p.
Katz, M., 1961. Acute toxicity of some organic insecticides to three
species of salmonids and to the threespine stickleback. Trans. Amer.
Fish. Soc., 90:264.
Korn, S. and R. Earnest, 1974. Acute toxicity of twenty insecticides
to striped bass, Morone saxatilis. Calif. Fish and Game, 60:128.
Lowe, J.I., 1966. Some effects of endrin on estuarine fishes. Proc.
19th Ann. Conf. S.E. Assn. Game and Fish. Commissioners, pp. 271-276,
Mount, D.I. and G.J. Putnicki, 1966. Summary report of the 1963
Mississippi fish kill. Trans. 31st No. Amcr. Wild!, and Nat. Res.
Conf., p. 177.
Post, G. and T.R. Schroeder, 1971. The toxicity of four insecticides to
four salmonid species. Bull. Envirdn. Contam. Toxicol., 6:144.
Schimmel, S.C., et. al_. 1975 Endrin: Effects on several estuarine
organisms. Proc. 29th Ann. Conf. S.E. Assn. Game and Fish. Commissioners,
Treon, et al_., 1955. Toxicity of endrin for laboratory animals. Jour.
Agric. and Food Chem., 3:842.
-------
6UTHION
CRITERION:
.01 ug/1 for freshwater and marine aquatic life.
RATIONALE:
Ninety-six-hour LC$Q values for fish exposed to the organophosphorus
pesticide, guthion, range from 4 to 4,270 ug/1 (Katz, 1961; Pickering, et al.,
1962; Lahav and Sarig, 1969; Macek, e£ 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 promelas,
exposed during a complete life cycle. An estimated "safe" long-term
exposure concentration for fathead minnows lies between 0.3 and 0.5 ug/1.
Survival of larvae was reduced at approximately 0.7 ug/1.
An investigation of the persistence of guthion in fish revealed that
50 percent of the chemical was lost in less than 1 week (Meyer, 1965).
Analysis of plankton and pondwater in the same study indicated a 50 percent
loss of guthion in about 48 hours. Flint, e_t al_. (1970) determined the
half-life of guthion at 30° C in pondwater and in a phosphate buffer
protected from light in the laboratory. The half-life in pondwater was
1.2 days whereas that in the laboratory solution was 10 days. The more
rapid degradation in pondwater was attributed to the effect of sunlight
and microorganisms..
Organophosphate pesticides are toxic because they inhibit the enzyme
acetylcholinesterase (AChE.) which is essential to nerve impulse conduction
2.71
-------
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, e_t aK, 1962; Weiss
and Gakstatter, 1964; Meyer, 1965). Weiss and Gakstatter (1964) found
that over a 15-day period bluegills, Lepomis macrochirus, exhibited AChE
inhibition at 1.0 ug/1 guthion but not at 0.1 ug/1. Exposure at 0.05 ug/1
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/1 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
I 111
discontinued (Weiss, 1959 and ; Carter, 1971). Inhibition of brain
AChE of fishes by 46 percent or more has been associated with harmful
effects in exposures to other 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_ aj_. ,In Press! In static tests, similar inhibition of AChE and mortality
were caused in the sheepshead minnow, Cyprinodon variegatus, in 2, 24, 48,
and 72 hours at concentrations of 50, 7, 3.5, and 3 ug/1, 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 20-
fold accumulation. LD$Q toxicity values for birds are relatively high and
range from 70 to 2,000 mg/kg (Tucker and Crabtree, 1970).
Ninety-six-hour LC^Q values for aquatic invertebrates range from
0.10 to 22.0 ug/1 (Nebeker and Gaufin, 1964; Gaufin, ejt al_., 1965; Jensen
and Gaufin, 1966; Sanders and Cope, 1968; Sanders, 1969 and 1972). Sanders
(1972) exposed the grass shrimp, Paleomonetes kadiakensi s, to guthion in
a continuous flow bioassay for up to 20 days and found that the 5-and
20-day 1659 values were 1.2 and 0.16 ug/1, respectively. He found that
the amphiood, Gammarus fasciatus, was the most sensitive aquatic organism
tested, with a 96-hour LC50 of 0.10 ug/1. 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 LC5Q values for Acroneuria pac.1f.1ca of 2.0 and 0.24 ug/1, respectively,
whereas for Pteronarcys californica the values were 4.6 and 1.3 ug/1,
respectively.
Results of other toxicity studies on marine organisms have
been reported. The 24-hour LC50 for the white mullet, Mugil curema, was
found to be 5.5 ug/1 guthion (Butler, 1963). The 96-hour LC5Q for the
striped mullet, Mugil cephalus, was determined by Lahav and Sarig (1969)
to be 8 ug/1 guthion. Portman (1972) reported the 48-hour 1X50 for tne
-------
fish, Pleuronectes limanda. to be 10 to 30 ug/1. The 48-hour LC5Q for
the European shrimp, Crangon crangon, was found to be 0.33 ug/1 guthion
(Portman, 1972). Butler (1963) found that the 24-hour EC5Q for blue crab,
Callinectes sapidus, was 550 ug/1 and the 48-hour EC50 for pink shrimp,
Penaeus duorarum, as 4.4 ug/1 guthion. The 48-hour TLm was estimated to
be 620 ug/1 for fertilized oyster eggs, Crassostrea virginica, and 860
ug/1 for fertilized clam eggs, Mercenaria mercenaria (Davis and Hidu, 1969),
A criterion level off.01 ug/1 for guthion is based upon use of
an 0.1 application factor applied to the 96-hour LC50 offcl ug/1 for
Gammarus and a similar value of 0.3 ug/1 for the European shriirp.
-------
REFERENCES CITED:
Adelman, I.R. and L.L. Smith. (Unpublished data.) Department of
Entomology, Fisheries and Wildlife, University of Minnesota, St. Paul.
Benke, 6.M. and S.D. Murphy, 1974. Anticholinesterase action of methyl
parathion, parathion and azinphosmethyl in mice and fish: Onset and
recovery of inhibition. Bull. Environ. Contam. Toxicol., 12:117.
Butler, P.A., 1963. Commercial fisheries investigations. In: pesticide-
wildlife studies during 1961 and 1962. U.S. Fish. Wild!. Serv. Circ.
167, Washington, D.C.
Carter, F.L., 1971. In vivo studies of brain acetylcholinesterase
inhibition by organophosphate and carbamate insecticides in fish.
Ph.D. Dissertation, Louisiana State University, Baton Rouge, Louisiana.
Coppage, D.L., 1972. Organophosphate pesticides: Specific level of
brain ACHE inhibition related to death in sheepshead minnows. Trans.
Amer. Fish. Soc., 101:534.
Coppage, D.L. and T.W. Duke, 1971. Effects of pesticides in estuaries
along the gulf and southeast Atlantic coasts. In: Proceedings of the
2nd Gulf Coast Conference on Mosquito Suppression and Wildlife Manage-
ment. (C.H. Schmidt, Ed.) National Mosquito Control - Fish and
Wildlife Management Coordinating Committee, Washington, D.C.
Coppage, D.L. and E. Matthews, 1974. Short-term effects of organo-
phosphate pesticides on cholinesterases of estuarine fishes and pink
shrimp. Bull. Environ. Contam. Toxicol., 11:483.
3*0
-------
Coppage, D.L., et.lL (In Press) Brain acetylcholinesterase inhibition in
fish as a diagnosis of environmental poisoning by malathion, 0,0-dimethyl
87(1,1-dicarbethoxyethyl) phosphorodithioate. Pesticide Biochemistry
and Physiology.
Davis, H.C. and H. Hidu, 1969. Effects of pesticides on embryonic develop-
ment of clams and oysters and on survival and growth of the larvae.
U.S. Fish and Wildlife Service, Fishery Bulletin, 67:393.
Eaton, J.G., 1970. Chronic malathion toxicity to the bluegill (Lepomis
macrochirus Rafinesque). Water Research, 4:673.
Flint, D.R., ertaK, 1970. Soil runoff, leaching, and adsorption and
water stability studies with guthion. Chemagro Rept. No. 28936.
Gaufin, A.R., e£al_., 1965. The toxicity of ten organic insecticides to
various aquatic invertebrates. Water and Sew. Works, 112:276.
Holland, H.T., et aK, 1967. Use of fish brain acetylcholinesterase to
monitor pollution by organophosphorus pesticides. Bull. Environ.
Contam. Toxicol., 2:156.
Jensen, L.D. and A.R. Gaufin, 1966. Acute and long-term effects of
organic insecticides on two species of stonefly naiads. Jour. Water
Poll. Cont. Fed., 38:1273.
Katz, M., 1961. Acute toxicity of some organic insecticides to three
species of salmonids and to the threespine stickleback. Trans. Amer.
Fish. Soc., 90:264.
-------
Lahav, M. and S. Sarig, 1969. Sensitivity of pond fish to cotnion
(azinphos methyl) and parathion. Bamidgeh, 21:67.
Macek, K.J., e_t al_., 1969. The effects of temperature on the susceptibility
of the bluegills and rainbow trout to selected pesticides. Bull. Environ.
Contam. Toxicol., 3:174.
Macek, K.J. and W.A. McAllister, 1970. Insecticide susceptibility of
some common fish family representatives. Trans. Amer. Fish. Soc., 99:20.
Meyer, P.P., 1965. The experimental use of guthion as a selective fish
eradicator. Trans. Amer. Fish. Soc., 94:203.
Nebeker, A.V. and A.R. Gaufin, 1964. Bioassays to determine pesticide
toxicity to the amphipod crustacean, Gammarus lacustris. Proc. Utah
Acad. Sc1. Arts and Letters, 41:64.
Pickering, Q.H., §t al_., 1962. The toxicity of organic phosphorus
insecticides to different species of warmwater fishes. Trans. Amer.
Fish. Soc., 91:175.
Portman, J., 1972. Results of acute tests with marine organisms, using
standard methods. In: Marine pollution and sea life (Ruivo, Ed.).
Fishing News (Books) Ltd, London, England, pp. 212-217.
Post, G. and R.A. Leasure, 1974. Sublethal effect of malathion to three
salmonid species. Bull. Environ. Contam. Toxicol., 12:312.
«•
Sanders, H.O., 1969. Toxicity of pesticides to the crustacean, Gammarus
lacustris. U.S. Dept. of the Interior, Bur. Sport Fish, and Wildl.,
Technical Paper No. 25, Washington, D.C.
-------
Sanders, H.O., 1972. Toxicity of some insecticides to four species of
malacostracan crustaceans. U.S. .Dept. of the Interior, Bur. Sport
Fish, and Wildl. Technical Paper No. 66, Washington, D.C.
Sanders, H.O. and O.B. Cope, 1968. The relative toxicities of several
pesticides to naiads of three species of stoneflies. Limnol. and
Oceanog., 13:112.
Tucker, R.K. and D.G. Crabtree, 1970. Handbook of toxicity of pesticides
to wildlife. U.S. Dept. of the Interior, Bur. Sport Fish, and Wildl.,
Resource Publ. No. 84, Washington, D.C.
Wei,ss, C.M., 1958. The determination of cholinesterase in the brain
tissue of three species of freshwater fish and its activation in vivo.
Ecology, 39:194.
Weiss, C.M., 1959. Response of fish to sub-lethal exposures of organic
phosphorus insecticides. Sew. and Ind. Wastes, 31:580.
Weiss, C.M., 1961. Physiological effects of organic phosphorus insecticides
on several species of fish. Trans. Amer. Fish. Soc., 90:143.
Weiss, C.M. and J.H. Gakstatter, 1964. Detection of pesticides in water
by biochemical assay. Jour. Water Poll. Cont. Fed., 36:240.
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HEPTACHLOR
CRITERION:
.001 ug/1 for freshwater and marine aquatic life.
The persistence, bioaccumulation potential and
carcinogenicity of heptachlor cautions human
exposure to a minimum.
RATIONALE;
The acute toxicity of heptachlor to mammals is generally low;
however, aquatic organisms exhibit sensitivity to this pesticide
at microgram per liter levels. Such levels range from a 96-hour
LCjjg of 3 ug/1 for the juvenile striped bass, Morone saxatilia
(Horn and Earnest, 1974) to a 96-hour LCso of 111.9 ug/1 for the
threespine stickleback, Gasterosteus aculeatus (Katz, 1961).
Sanders and Cope (1968) reported even lower 96-hour LC50 values for
three species of stoneflies; Pteronarcys californica, 1.1 ug/1;
Pteronarcella badia, 0.9 ug/1 and Claassenia sabulosa, 2.8 ug/1.
Anderson (1960) found that Daphnia magna were immobilized
after 50 hours exposure to 57.77 ug/1. A 48-hour LC5Q of 42 ug/1
heptachlor for Daphnia pulex was reported by Cope (1966) .
Sanders and Cope (1968) determined the 96-hour LC50 values for the
stoneflies, Pteronarcys californica, Pteronarcella badia and
Claassenia sabulosa to be 1.1, 0.9 and 2.8 ug/1, respectively.
-------
Sanders (1972) found the grass shrimp, Palaeroonetes kadiakensia,
with a 96-hour l£$Q of 1.8 ug/1, to be the most sensitive among
three crustaceans tested in static bioassays. Stoneflies,
therefore, appear to be the most sensitive group of freshwater
organisms among those tested.
Cope (1966) found the 48-hour LC50 for heptachlor to the
bluegill, Lepomis macrochirus, and rainbow trout, Salmo
gairdneri, to be 26 ug/1 and 9 ug/1, respectively. The 96-hour
LCsO of heptachlor for the bluegill was determined by Weiss (1964)
to be 19 ug/1. Katz (1961) found that the 96-hour LCSO's of
heptachlor for the coho salmon, Oncorhynchus kisutch; chinook
salmon, Oncorhynchus tschawytscha; rainbow trout, Salmo
gairdneri; and.threespine stickleback, Gasterosteus aculeatus,
were 59, 17.3, 19.4, and 111.9 ug/1, respectively. The 96-hour
^50 of heptachlor to the fathead minnow, Pimephales promelas;
bluegill, Lepomis macrochirus; goldfish, Carassius auratus; and
guppy, Poecilia reticulata, were determined to be 94, 230 and 170
ug/1, respectively (Henderson, et al., 1959).
Data are available on marine animals exposed to heptachlor in
96-hour flow-through bioassays. The 96-hour LC50 of the marine
bluehead, Thalassoma bifasciatum, was determined to be 0.8 ug/1
(Eisler, 1970). Korn and Earnest (1974) determined the 96-hour
LC50 using juvenile striped bass, Morone saxati 1 is ,^to be 3.0
ug/1. Schimrnel e_t al_., (In Press) reported the following 96-hour
LCBO's in flow-through bioassays: pink shrimp, Penaeus duorarum
0.11 ug/1; grass shrimp. .Palaemonetes vulgaris ^1.06 ug/1;
sheepshead minnow, Cyprinodon variegatus, 3.68 ug/1; pinfish,
Lagodon rhomboides 3.77 ug/1; spot. Leiostomus xanthurus. V0.85 ug/1.
Shell deposition of the American oyster. Crassostrea virgim'ca twas
inhibited by 50 percent (EC50) at 1.5 ug/1 heptachlor.
-------
Heptachlor will accumulate in the food chain.
Wilson (1965) demonstrated that oysters can concentrate heptachlor
almost 18,000 times (wet weight basis). Andrews et al.
(1966) reported concentration factors as high as 1,840 in field
tests with bluegill, Lepomis macrochirus. Since the effective
(as opposed to the applied) concentration was apparently lower
than the initial levels because of sorption and biological
uptake, true concentration factors must have been higher than
reported. For example, 24 hours after treatment with 24 ug/1 the
pond water contained only 2 ug/1. Using this concentration and
the 46.0 ppm residue found in the fish, a concentration factor of
23,000 can be calculated. The involvement of several trophic
levels might also increase the degree of accumulation.
Schimmel, e_t al_., (In Press) reported that heptachlor concentration factors
ranged from 2,800 to 21,300 in estuarine fishes exposed for 96 hours;
oyster bioconcentration factors ranged from 4,500 to 8,500, arid factors of
206 to 700 occurred in similar tests with crustaceans.
Birds are sensitive to low dosages of heptachlor in their
diet. Heath, e£ al. (1972) found the 5-day LC for the young of
four species to range from 24 to 54 ppm. All woodcock died from
a dietary dosage of 0.72 ppm and some died from 0.22 ppm (Stickel
and Stickel, 1965). Residues of this magnitude (0.2 ppm) would
result in fish or other aquatic life if they were exposed to 0.01
ug/1 of heptachlor and accumulated it at 20,000 times water
-------
concentrations. Such residues would pose a hazard to birds as
sensitive as woodcock that fed on these fish. Residues of this
magnitude fall just under the 0.3 ppm guideline established by
the U.S. Food and Drug Administration as the limit allowed in
edible fish tissue (U.S. Food and Drug Administration, 1974).
Since heptachlor is a highly persistent chemical which
bioaccumulates in aquatic organims used for human food and also
is considered a potential human carcinogen (Train, 1974) levels
of heptachlor in waterways should be kept as low as feasible.
The July 1975 action by the Environmental Protection Agency in
suspending the production and use of heptachlor should result in
a gradual decrease in concentrations in the environment. Any
addition of heptachlor to water should be considered potentially
hazardous to humans. Such additions should not be permitted
without substantial documentation that alternatives are either
infeasible or potentially more hazardous. The persistence,
bioaccumulative properties and carcinogenic potential of
heptachlor should be taken into account when determining the uses
of water containing measurable amounts of heptachlor.
Using an 0.01 application factor to the 96-hour LC50
of 0.1 ug/1 for the pink shrimp, a marine criterion of 0.001
ug/1 is obtained. Toxicity and demonstrated bioaccumulation
potential of heptachlor in freshwater biota reflects the
applicability of the marine criterion to freshwater.
-------
REFERENCES CITED:
Anderson, B. G., 1960. The toxicity of organic insecticides
to Daphnia. In: Trans. 1959 Seminar Biol. Probs. Water
Pollution. Robt. A. Taft San. Center, Tech. Dept. W60-
3, p. 94.
Andrews, A.K., et al., 1966. Some effects of heptachlor on
bluegills (Lepomis macrochirus) . Trans. Amer. Fish.
Soc., 95:297.
Cope, O.B., 1966. Contamination of the freshwater ecosystem
by pesticides. Jour. Appl. Ecol. (Suppl.), 33.
Eisler, R., 1970. Acute toxicity of organochlorine and
organophosphorus insecticides to estuarine fishes.
Tech. Papers, Bur. Sport Fish. Wildl., 46.
Food and Drug Administration, 1974. Poisonous or deleterious
substances in peanuts, evaporated milk, fish and
shellfish. Proposed Rules. Federal Register, December
6, 1974. Washington, D. C.
-------
Heath, R. G., et a]L., 1972. Comparative dietary toxicities
of pesticides to birds. Bureau of Sport Fisheries and
Wildlife, Wildlife Report No. 152, U.S. Dept. of the
Interior, Washington, D. C.
Henderson, D., et al., 1959. Relative toxicity of ten
chlorinated hydrocarbon insecticides to four species of
fish. Trans. Amer. Fish. Soc., 90: 264.
Katz, M., 1961. Acute toxicity of some organic insecticides
to three species of salmonids and to the threespine
stickleback. Trans. Amer. Fish. Soc., 90: 264.
Horn, S., and R. Earnest, 1974. Acute toxicity of twenty
insecticides to striped bass, Morone saxatilis. Calif.
Fish and Game, 60:128.
Sanders, H.O., 1972. Toxicity of some pesticides to four
species of malocostracan Crustacea. U.S. Bureau of
Sport Fisheries and Wildlife Technical Paper No. 66.
Sanders, H.O. and O.B. Cope, 1968. The relative toxicities
of several pesticides to naiads of three species of
stoneflies. Limnol. and Oceanog., 13:112.
Schimmel, S.C., et al., 1976. _ Heptachlor: Toxicity
to and uptake by several estuarine organisms. Jour. Toxlcol.
& Environ. Health (In Press).
-------
Stickel, W.H. and L.F. Stickel. 1965. Effects of hsptachlor-
contaminated earthworms on woodcocks, jour. Wildlife
Management, 21:132.
Train, R. E., 1974. Pesticide products containing heptachlor
or chlordane: Intent to cancel registrations, Federal
Register, Vol. 39, No. 229, Tues., Nov. 26, 1974, p
41299.
Weiss, C.M., 1964. Organic pesticides and water'pollution.
Public Works 95:84.
Wilson, A.J., 1965. Chemical assays. In: Annual Report of
the Bureau of Commarical Fisheries Biological
Laboratory. Gulf Breeze, Florida, Fiscal Year 1965. U.
S. Bur. Comm. Fish. Circ. 247.
-------
LINDANE
CRITERIA:
4.0 ug/1 for domestic water supply (health);
0.01 ug/1 for freshwater aquatic life;
0.004 ug/1 *or marine aquatic li^e
RATIONALE;
The highest level of lindane found to have minimal or no long-term
effects in the most sensitive .-mammal tested, the dog, is 15.0 mg/kg in
the diet or 0.3 mg/kg of body weight/day (Lehman, 1965). Where adequate
human data are not available for corroboration of the animal results,
the total "safe" drinking intake level is assumed to be 1/500 of the "no
effect" or "minimal effect" level reported for the most sensitive
animal tested.
Applying the available data and based upon the assumption that
20 percent of the total intake of lindane is from drinking water, that
the average person weighs 70 kg and consumes 2 liters of water per day,
the formula for calculating a criterion is 0.3 mg/kg x 0.2 x 70 kg x 1/500 x
1/2 * .004 mg, thus deriving the criterion level for domestic water
supply of 4 ug/1.
The brown trout, Salmo trutta, apparently is the fish most sensitive
to lindane among those species on which aquatic bioassays have been
performed, with a 96-hour LC50 of 2 ug/1 (Macek and McAllister, 1970).
Several authors (Boyd and Ferguson, 1964; Naqvi, et aj_., 1969;
Minchew and Ferguson, 1970) have documented increased
resistance to lindane toxicity among fish an>1
-------
invertebrates experiencing previous exposure to the chemical. The most
sensitive invertebrates tested appear to be only slightly more sensitive
than fish. Two investigators (Snow, 1958; Cope, 1965) reported TLm
values of 1 ug/1 for stoneflies. Sanders and Cope (1968) reported a
TLra for stoneflies of 4.5 ug/1 lindane. Macek, et aJL (1974) determined
the acute and chronic toxicities of lindane to Daphnia magna; the midge,
Chironomus tentans; and the scud, Gammarus fasciatus. The midge was the
most sensitive of these species, with 2.2 ug/1 being the highest concen-
tration producing no observable adverse effect. A concentration of 11
ug/1 was determined to be'"safe" for Daphnia. the least sensitive, over
three consecutive generations of exposure.
A criterion of 0. 01 ug/1 for fresh waters is derived by applying
an application factor of 0. 01 to the TLM for the stonefly. 1.0 ug/1.
The brown trout, a sensitive human food organism, was shown
to have a 96-hour LC of 2.0 ug/1. Hence the criterion level
50
would provide for a margin of safety for both recreationally important organisms
as well as trophic levels in the aquatic environment which are significant in the food
chain.
Only limited information is available on accumulation of lindane
in fish tissues. However, Macek, et al_. (1974) observed whole-body
(eviscerated) levels (wet weight) of 500 times the corresponding water
concentrations in fathead minnows, Pimephales promglas, that had been
exposed for several months. Butler (1967) observed accumulations of up
to 250 tunes exposure concentrations in marine nollusks (wet weight basis).
-------
The recommended guideline of the U.S. Food and Drug Administration
for lindane in edible fish tissue is 0.3 mg/kg (FDA, 1974). Thus, if
the observed 500-fold accumulation were to occur, a lindane criterion
of 0.01 ug/1 in water would result in a tissue concentration in freshwater
fish of .005 mg/kg, which is well below the FDA guideline.
Eisler (1969) reported 96-hour LC$Q values of 5.0 ug/1 for both the
hermit crab, Pagurus longi carpus, and the sand shrimp, Crangon
septemspinosa. and 10 ug/1 for the grass shrimp, Palaemonetes pugio.
Butler (1963) found the following 48-hour LC$Q values: brown shrimp,
Penaeus aztecus - 0.4 ug/1; juvenile white mullet, Mugil curema -
30 ug/1; and the longnose k.illifish, Fundulus similis - 240 ug/1. A
96-hour LC5o of 7.3 ug/1 for the striped bass, Morone saxatilis was
reported by Korn and Earnest (1974). schimmel . (unpublished data)
found the pink shrimp, Penaeus duorarum, 96-hour IC50 to be 0.17 ug/1,
grass shrimp Palaemonetes pugio, 4.4 ug/1 and pinfish, Laaodon rhomboidesr
30.6 ug/1 in flow-through bioassays.
Until additional data on the effect of lindane to marine organisms
, it is reconmended that a marine criterion be based on 0.01
of the 0.4 ug/1 96-hour 1C value for the brown shrimp. The marine
50
criterion is- therefore, 0-004 ug/1.
Unpublished data suggest that the pink shrimp is a more sensitive
species than the brown shrimp; however for oumoses of achieving
a criterion the published data were used. A level of 0.004 ug/1 should
provide a nargin of safety for both sensitive species.
-------
REFERENCES CITED:
Boyd, C.E. and D.E. Ferguson, 1964. Susceptibility and resistance
of mosquitofish to several insecticides. Econ. Ehtomol., 57:430.
Butler, P.A., 1963. Commercial fisheries investigations. In: Pesticide
wildlife studies. U.S. Dept. of Interior, Fish and Wildlife Service,
Washington, D. C., Circ. 167, p. 11-25.
Butler, P.A., 1967. Pesticide residues in estuarine mollusks. Ijn:
P.F. McCarty and R. Kennedy. Proceedings of the National Symposium
on Estuarine Pollution. Stanford, California, Stanford University,
Department of Civil Eng., pp. 107-121.
Cope, O.B., 1965. Sport fishery investigations. In: Effects of pesticides
on fish and wildlife, 1964. Research findings of the Fish and Wildlife
Service, U.S. Fish and Wildlife Service, Circ. 226. pp. 51-63.
Eisler, R., 1969. Acute toxicities of insecticides to marine decapod
crustaceans. Crustaceana, 16:302.
Food and Drug Administration, 1974. Poisonous or deleterious substances
in peanuts, evaporated milk, fish and shellfish. Proposed Rules.
Federal Register. December 6, 1974, Washington, D. C.
Korn, S. and R. Earnest, 1974. Acute toxicity of twenty insecticides
to striped bass, Morone saxatilis. Calif. Fish and Game, 60:128.
-------
Telman, A.-T. , 1965. Sumnaries of pesticide toxicity. Assoc.
of Food and Drug Officials of the U.S., Topeka, Kansas, pp 1-40.
Macek, K.J. and McAllister, 1970. Insecticide susceptibility
of sane connon fish family representatives. Trans. Amer. Fish
Sec., 99:20.
1
Macek, K.J., et, aK, 1974. Chronic toxicity of lindane to selected
aquatic invertebrates and fishes. Environmental Protection Agency
Research Contract Report, Environmental Protection Agency Ecol.
Res. Series, (In preparation).
Minchew, G.D. and D.E. Ferguson, 1970. Toxicities of six insecticides
to resistant and susceptible green sunfish and golden shiners in static
bioassays. Jour. Miss. Acad. Sci., 15:29.
Naqvi, S.M. and D.E. Ferguson, 1969. Pesticide tolerances of selected
freshwater invertebrates. Jour. Miss. Acad. Scl., 14:121.
Sanders, H.O., and O.B. Cope, 1968. The relative toxicities of several
pesticides to naiads of three species of stoneflies. Limnol. and
Oceanog., 13:112.
Scniimel, S.C., 1976. Semi-Annual Report' Gulf Breeze Environmental
Research Laboratory (In Press).
Snow, J.R., 1958. A preliminary report on the comparative testing of
some of the newer herbicides. Prcc. llth Ann. Conf., Southeast Assoc.
Game Fish Comm., pp. 125-132.
19*
-------
MALATHION
CRITERION;
0.1 ug/1 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 LC5Q
between 120 and 265 ug/1 for 4 species of salmonids. Macek and
McAllister (1970) found a 96-hour LC5Q range between 101 and 285
ug/1 for 3 species of centrarchids and 3 species of salmonids.
Other 96-hour LCSO's are: rainbow trout, Salmo gairdneri, 68 ug/1
(Cope, 1965); largemouth bass, Micropterus salmpides, 50 ug/1
(Pickering, et al», 1962) and chinook salmon, Oncorhynchus
tshawytscha, 23 ug/1 (Katz, 1961). All of the above tests were
in static systems. Eaton (1970) determined a 96-hour LC5Q for
bluegill, Lepomis macrochirus, in a flow-through system at 110
ug/1. Macek and McAllister (1970) reported a similar 96-hour 1^59
for the bluegill in a static exposure. Static 96-hour LCSOs of
120 and 160 ug/1 were reported by Post and Schroeder (1971) for brook trout,
Salvelinus fontinalis. Bender (1969) indicated that the acute toxicity to fathead
minnows, Pimephales promelas, is slightly greater (about 2.0
times) in a static system than in a flow-through system. The
flow-through acute 'toxicity to fathead minnows reported by Mount
and Stephan (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 LCso for Gammarus lacustris was 1.0
ug/1 (Sanders, 1969); for Pteronarcella badia, 1.1 ug/1 (Sanders
and Cope, 1968); and for Gammarus fasciatus, 0.76 ug/1 (Sanders,
1972). The 48-hour LC^Q for Simocephalus serrulatus was 3.5 ug/1
and for Daphnia pulex, 1.8 ug/1 (Sanders and Cope, 1966).
Daphnia were immobilized in 50 hours in 0.9 ug/1 (Anderson,
1960}. The 24-hour LCSOs for two species of midge larvae were
2.1 ug/1 (Mulla and Khasawinah, 1969) and 2.0 ug/1 (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 roost 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 one to
two 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/1
malathion in flowing seawater kills 40 to 60 percent of the
marine fish, Lagodon rhomboides, in 3.5 hours and causes about 75
percent brain acetylcholinesterase (AChE) inhibition (Coppage, et
al.t 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/1, respectively. A concentration of 31 ug/1
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,
Mucp.1 cephalus, in an estuary sprayed with malathion (3 oz./acre)
during a large-scale mosquito control operation had about 98
percent inhibition of brain AChE This is in agreement with 70
to 80 percent or more inhibition of brain AChE levels at and
below which some deaths are likely to occur in short-term
exposure. Spot, Leiostomus xanthurus, and Atlantic croaker,
Micropogon undulatus, also had substantial inhibition of brain
AChE 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 seve
fishes at 20° C in static, aerated seawater. The 96-hour
values (in ug/1) were: Menidia menidia, 125; Mugil gephalus, 550;
Fundulus majalis, 250; Fundulus heteroclitus, 240; Sphaeroides
maculatus, 3,250; Anguilla rostrata, 82, and Thalassoma
bifasciatum, 27. Katz (1961) reported the static 24 hour LC50 for
Gasterosteus aculeatus in 25 o/oo saltwater as 76.9 ug/1 active
ingredient. The 96-hour LC^Q for striped bass, Morone saxatilis,
in intermittent flowing seawater has been reported as 14 ug/1
(U.S. BSFW 1970).
Reporting on studies of the toxicity of malathion on marine
invertebrates, Eisler (1969) found the 96-hour LCso (static, 24
o/oo salinity aerated) to be 33 ug/1 for sand shrimp, Crangon
septemspinosa; 82 ug/1 for grass shrimp, Palaemonetes vulgaris;
.and 83 ug/1 for hermit crab, Pagurus longicarpus. Growth of
oyster, Crassostrea virginica, was reduced 32 percent by 96-hour
exposure to 1 mg/1 (Butler, 1963). The 48-hour LC5Q for fertilized
eggs of oysters was estimated by Davis and Hidu (1969) to be 9.07
mg/1 and the 14-day LC50 for larvae, 2.66 mg/1.
Malathion enters the aquatic environment primarily as a
result of its application as an insecticide. Because it degrades
quite rapidly inrmost waters depending on pH, its occurrence
is sporadic rather than continuous. Because the toxicity
-------
1s exerted through inhibition of the enzyme acetylcholinesterase (AChE)
and because such inhibition may be additive with repeated exposures and
may be caused by any of the organo-phosphorus insecticides, inhibition of
AChE by more than 35 percent may be expected to result in damage to
aquatic organisms.
An application factor of 0.1 is applied to the 96-hour LC50 data
for Gammarus lacustris, G. fasciatis and Daphnia , which are all
approximately 1,0 ug/1, yielding a criterion of 0.1 ug/1.
-------
REFERENCES CITEDj
Anderson, B.C., 1960. The toxicity of organic insecticides
to paphnia. Second Seminar on Biol. Problems in Water
Pollution. Robert A. Taft Sanitary Engineering Center
Technical Report W60-3, Cincinnati, Ohio.
Bender, M.E., 1969. The toxicity of the hydrolysis and
breakdown products of malathion to the fathead minnow
(Pimephales prgmelas, Rafinesgue). Water Res., 3:571.
Butler, P.A., 1963. Commercial fisheries investigations.
In: Pesticide Wildlife Studies during 1961 and 1962.
U.S. Fish Wildl. Ser. Circ. 167, Washington, D.C.
Coppage, D.L. and T.W. Duke, 1971. Effects of pesticides in
estuaries along the Gulf and Southeast Atlantic Coasts.
In: Proceedings of the 2nd Gulf Coast Conference on
Mosquito Suppression and Wildlife Management. (C.H.
Schmidt, ed.) National Mosquito Control - Fish and
Wildlife Management Coordinating Committee, Washington,
D.C.
-------
Coppage, D.L. and E. Matthews, 1974. Short-term effects of
organophosphate pesticides on cholinesterases of
estuarine fishes and pink shrimp. Bull. Environ.
Contam. Toxicol. lit 483.
Coppage, D.L., et al., 1975. Brain acetylcholinesterase
inhibition in fish as a diagnosis of environmental
poisoning by malathion, £, 0-dimethyl £-(1,
1-dicarbethoxy-ethyl) phosphorodithioate. Pesticide
Biochemistry and Physiology (in press).
Cope, O.B., 1965. Sport fishery investigations. Ini The
Effects of Pesticides on Fish and Wildlife. U.S.
Department of the Interior, Washington, D.C. Fish and
Wildlife Service Circular 266.
Davis, H.C. and H. Hidu, 1969. Effects of pesticides on
embryonic development of clams and oysters and on
survival and growth of the larvae. U.S. Fish and
Wildlife Service, Fishery Bulletin. 67: 393.
Eaton, J.G., 1970. Chronic malathion toxicity to the
bluegill (Lepomis macrochirus Rafinesque). Water Res.,
4t 673.
302,
-------
Eichelberger, J.W. and J.J. Lichtenberg, 1971. Persistence
of pesticides in river water. Environ. Sci. & Technol.,
5: 541.
Eisler, R., 1969. Acute toxicities of insecticides to marine
decapod crustaceans. Crustaceana. 16: 302.
Eisler, R., 1970. Acute toxicities of organochlorine and
organophosphorus insecticides to estuarine fishes. U.S.
Bureau of Sport Fisheries and Wildlife, Technical Paper
46.
Henderson, C. and Q.H. Pickering, 1958. Toxicity of organic
phosphorus insecticides to fish. Trans. Amer. Fish.
Soc., 87: 39.
Karnak, R.E. and W.J. Collins, 1974. The susceptibility to
selected insecticides and acetylchlolinesterase activity
in a laboratory colony of midge larvae, Chironomus
tentans (Diptera: chironomidae). Bull. Environ.
Contain. Toxicol., 12: 62.
Katz, M., 1961. Acute toxicity of some organic insecticides
to three species of salmonids and to the threespine
stickleback. Trans. Araer. Fish Soc. 90: 264.
-------
Macek, K.J. and W.A. McAllister, 1970. Insecticides
susceptibility of some common fish family
representatives. Trans. Amer. Fish. Soc., 99: 20.
Mount, D.T. and C.E. Stephan, 1967. A method for
establishing acceptable toxicant limits for
fish-malathion and the butoxyethanol ester of 2,4-D.
Trans. Amer. Fish. Soc., 21: 185.
Mulla, M.S. and A.M. Khasawina, 1969. Laboratory and field
evaluations of larvicides against chironomid midges.
Jour. Econ. Entomol., 62: 37.
Paris, D.F., et al_ 1975.Rates of degradation of raalathion by
bacteria isolated from aquatic systems. Environ. Sci. &
Technol., 9: 135.
Pickering, Q.H., et al., 1962. The toxicity of organic
phosphorus insecticides to different species of
warmwater fishes. Trans. Araer. Fish. Soc., 91: 175.
Post, G. and T.R. Schroeder, 1971. The toxicity of four
insecticides to four salmonid species. Bull. Environ.
contam. Toxicol., 6: 144.
-------
Sanders, H.O., 1969. Toxicity of pesticides to the
crustacean, Gammarus lacustris. U.S. Department of the
Interior, Washington, O.C., Bureau of Sport Fisheries
and Wildlife Technical Paper No. 25.
Sanders, H.O., 1972. Toxicity of some insecticides to four
species of malacostracan crustaceans. U.S. Department
of the Interior, Washington, D.C. Bureau of Sport
Fisheries and Wildlife Technical Paper No. 66.
Sanders, H.O. and O.B. Cope, 1966. Toxicities of several
pesticides to two species of cladocerans. Trans. Amer.
Fish. Soc., 95: 165.
Sanders, H.O. and O.B. Cope, 1968. The relative toxicities
of several pesticides to naiads of three species of
stonefly. Limnol. & Oceanog., 13: 112.
U.S. Bureau of Sport Fisheries and Wildlife, 1970. Resource
Publication 106. Washington, D.C.
Weiss, C.M. and J.H. Gakstatter, 1964. The decay of
anticholinesterase of organic phosphorus insecticides on
storage in waters of different pH. Advances in Water
Pollution Research, 1: 83.
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METHOXYCHLOR
CRITERIA;
100 ug/1 for domestic water supply (health);
0.03 ug/1 for freshwater and marine aquatic life.
RATIONALE:
The highest level of raethoxychlor found to have minimal or no
long-term effects in man is 2.0 mg/kg of body weight/day (Lehman,
1965). Where adequate human data are available for corroboration-
of the animal results, the total "safe" drinking water intake
level is assumed to be 1/100 of the "no effect" or "minimal
effect" level reported for the most sensitive animal tested, in
this case, man.
Applying the available data and based upon the assumptions
that 20 percent of the total intake of methoxychlor is from
drinking water, 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/1.
A criterion level for domestic water supply of 100 ug/1 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 prcmelas,
embryos at 0.125 ug/1 and lack of spawning at 2.0 ug/1. Yellow perch,
Perca flavescens, exposed to 0.6 ug/1 for eiaht months exhibited
reduced growth. The 96-hour LC5Q concentration was 7.5 and 22 ug/1
for the fathead minnow and yellow perch, respectively. Korn and Earnest
(1974) obtained a 96-hour LC of 3.3 ug/1 with juvenile striped bass,
Morone saxatilis, exposed to methoxychlor in a flowing-^water bioassay.
Sanders (1972) determined a 96-hour LC50 value of 0.5 ug/1 for
the crayfish, Orconectes nais. Merna and Eisele (1973) obtained
a 96-hour LCso value of 0.61 ug/1 for the scud, Gammarus
pseudolimnaeus and 96-hour LC50s ranging from 1.59 to 7.05 ug/1
for the crayfish, Orconectes nais, and three aquatic insect
larvae. In 28-day exposures, reduction in emergence of mayflies,
Stenonema sp., and in pupation of caddisflies, Cheumatopsyche
S£., were observed at 0.5 and 0.25 ug/1 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/1 methoxychlor for one year.
«•
The near extinction of one species of scud, Hyallella azteca, and
30?
-------
reductions in populations of other sensitive species, as well as
biomass were observed. Residue accumulation of up to 1,000 times
the level in the stream was observed in first-year crayfish,
Orconectes nais. Metcalf, et al. (1971) traced the rapid
conversion of methoxychlor to water soluble compounds and
elimination from the tissues of snails, mosquito larvae, and
mosquitofish. Thus, methoxychlor appears to be considerably less
bioaccumlative in aquatic organisms than some of the other
chlorinated pesticides.
Methoxychlor has a very low accumulation rate in birds and
mammals (Stickel, 1973), and relatively low avian (Heath, et al.,
1972), and mammalian (Hodge, et al., 1950) toxicities. No
administrative guidelines for acceptable levels in edible fish
tissues have been established by the U.S. Food and Drug
Administration.
The above data indicate that 0.1 ug/1 methoxychlor 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 0.03 ug/1 is recommended. This
criterion should protect fish as sensitive as striped bass and is
ten times lower than the level causing effects on some
invertebrate populations in a one-year dosing of a natural
stream.
-------
Bahner and Niratno (1974) found the 96-hour LC50 of methpxychlor for the
pink shrimp/ Penaeus duorarum, to be 3.5 ug/Jl and the 30-day LC50 to be
1.3 ug/1. Using an application factor of 0.01 with the pink shrinp acute
toxicity of 3.5 ug/1/ the recommended criterion for the marine environment
is 0.03 ug/1.
Butler (1971) found accumulation factors of 470 and 1,500 for
.the molluscs, Mercenaria mercenaria, and Mya arenaria,
respectively, when exposed to 1 ug/1 methoxychlor for 5 days.
i
Using the 1,500. accumulation factor as a basis, a water
concentration of 0.2 ug/1 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/1 is
an order of magnitude lower than this concentration.
-------
CITED;
Bahner, L.H. and D.R. Niitroo, 1974. Methods to assess effects of
combinations of toxicants, salinity, and temperature on estuarine
animals. Proc. 9th Am. Confr. on Trace Substances :Ln Env. Health,
Univ. Miss. Columbia, Mo.
Butler, P.A. 1971. Influence of pesticides on marine
ecosystems. Proc. Royal Soc. Lond., 177:321.
^Lsele, P.J. , 1974. The effects of nethoxychlor on aquatic
invertebrate populations and coramunities, Ph.D. Thesis,
Univ. of Michigan, Ann Arbor, 151 p.
Heath, R.G. , et al. , 1972. Comparative dietary toxicities of
pesticides to birds. Bureau of Sport Fisheries and
Wildlife, Wildlife Report No. 152, U.S. Dept. of the
Interior, Washington, D.C., 57 p.
Hodge, B.C., et al., 1950. Short-term oral toxicity tests of
methoxychlor in rats and dogs. Jour. Pharmacol. Exp.
Therapy, 99:140.
Kern, S. and R. Earnest, 1974. Acute toxicity of twenty
insecticides to striped bass, Morone saxatilis. Calif.
Fish and Game, 60:128.
Lehman, A. J. , 1965. Summaries of pesticide toxicity. Assoc.
of Food and Drug Officials of the U.S., Topeka, Kansas,
pp. 1-40.
3/0
-------
Merna, J.W. and P.J. Eisele, 1973. The effects of
methoxychlor on aquatic biota. U.S. EPA Ecological Res.
Series, No. EPA- R3-73-046. U.S. Government Printing
Office, Washington, D.C.
Metcalf, R.L., et al., 1971. Biodegradable analogs of DDT.
Bull. World Health Organization, 44:363.
Sanders, H«O., 1972. Toxicity of some insecticides to four
species of malacostracan crustaceans. Tech. Paper No.
66, Bureau of Sport Fisheries and Wildlife. U.S.
Government Printing Office, Washington, D.C.
Stickel, L.F., 1973. Pesticide residues in birds and
mammals. In: Environmental pollution by pesticides.
(C.A. Edwards, Ed.) Plenum Press, New York, pp. 254-312,
111
-------
MIREX
CRITERION;
0.001 ug/1 for freshwater and marine aquatic life.
RATIONALE;
Mirex is used to control the imported fire ant Solenopsis saevissima
richteri in the southeastern United States. Its use is essentially limited
to the control of this insect and it is always presented in bait. In the
most common formulation, technical grade mirex is dissolved in soybean
oil and sprayed on corncob grits. The bait produced in this manner
consists of 0. 3 percent mirex, 14. 7 percent soybean oil, and 85 percent
corncob grits. The mirex bait often is applied at a rate of 1. 4 kilograms
per hectare, equivalent to 4. 2 grams of toxicant per hectare.
Relatively few studies have been made of the effects of mirex on
freshwater invertebrates. Of these, only Ludke, etal. (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/1 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/1 for 144
hours were dead. Exposure to 5 ug/1 for 6, 24, and 58 hours resulted
in 26, 50, and 98 percent mortality 10 days after transfer to clean water.
-------
Crayfish, Procambarus hayi, were exposed to 0.1 and 0. 5 ug/1 for
48 hours. Four days after transfer to clean water, 65 percent of the
animals exposed to 0.1 ug/1 were dead. At the 0. 5 ug/1 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 concen-
trations of 0.86 ug/1. 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 two bait particles each.
From this report it is obvious that mirex is extremely toxic to these
species of crayfish. Mortality andaccumulation increases with time of
exposure to the insecticide. Concentrations_as low as 0.1 uq/1 or the
inqestion of one Darticle 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 bait per hectare) to four ponds containing channel catfish,
Ictalurus punctatus. Three applications were made over an 8-month
period with the first application 8 days after fingerling (average weight
18. 4 g) catfish were placed in the ponds. Fish were collected at each
subsequent application (approximately 4-month intervals). Two and one-
half months after the final application, the ponds were drained, all fish
ware measured, weighed, and the percent survival was calculated.
-------
Mi rex 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_. (1968) exposed bluegills, Lepomis
macrochirus. and the goldfish, Carassius auratus, to mirex by feeding a
mirex-treated diet (1, 3, and 5 mg mirex per kg body weight) or by treat-
ing holding ponds with mirex bait (1. 3, 100, and 1000 ug/1 computed water
concentration). They reported no mortality or tissue pathology for the
bluegills; however, after 56 days of exposure, gill breakdown in goldfish
was found in the 100 and 1000 ug/1 contact exposure ponds, and kidney
breakdown was occurring in the 1000 ug/1 ponds. Mortality in the feeding
experiments was not related to the level of exposure, although growth of
the bluegills fed 5 ug/1 mirex was reduced.
In laboratory and field test systems reported concentrations of mirex
usually are between 0. 5 and 1. 0 ug/1 (Van Valin, et^aL, 1968; Ludke,
et al., 1971). Although mirex seldom is found above 1 ug/1 in the aquatic
environment, several field studies have shown that the insecticide is
accumulated through the food chain. Borthwick, et^al. (1973) reported
the accumulation of mirex in South Carolina estuaries. Their data revealed
that mirex was transported from treated land and marsh to the estuary
animals and that accumulation, especially in predators, occurred. In the
test area, water samples consistently were less than 0.01 ug/1. 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
3/v
-------
and 78 percent of the birds contained residues up to 17 ug/g. Naqvi
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 environ-
ment. Wolfe and Norment (1973) sampled an area for one year following
an aerial application of mirex bait (2.1 g mirex/hectare). Crayfish residues
ranged from 0. 04 to 0.16 ug/g. Fish residues were about 2 to 20 times
greater than the controls and averaged from 0. 01 to 0. 76 ug/g. Kaiser (1974)
reported the presence of Mirex in fish from the Bay of Quints, Lake
Ontario, Canada. Concentrations range from 0.02 ug/g in the gonads of the
aorthern long nose gar, Lepistosteus psseus, to 0.05 ug/g in the areal
fin of northern pike, Esox lucius. 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 (Stickel, e^aL , 1973) that the persistence
of mirex in bird tissue exceeds that of all organochlorine compounds
tested except for DDE. Delayed mortality occurred among birds sub-
jected 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. Bio accumulation
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/1 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/1 technical grade
mirex in flowing sea water was lethal to juvenile pink shrimp, Penaeus
durorarum, in a three-week exposure (Lowe, e£ al., 1971). In static tests
with larval stages (megalopal) of the mud crab, Rhithropanopeus harrisii,
reduced survival was observed in 0.1 ug/1 mirex (Bookhout, et al., 1972).
MJB1). In three of four 28-day seasonal flow-through experiments, Tagatz, et al.,
(1975) found reduced survival of Callinectes sapidus, Penaeus durorarnm. and
grass shrimp, Palaemonetes pugio, at levels of 0.12 ug/1 in summer, 0.06 ug/1
in fall, and 0.09 ug/1 in winter.
Since two reports, Lowe, et al., (1971) and Bookhout, e£al_., (1972),
reported that effects of mirex on estuarine and marine crustaceans were
observed only after considerable time had elapsed, it seems reasonable
that length of exposure is an important consideration for this chemical.
This may not be the case in fresh water since the crayfish were affected
within 48 hours. Therefore, a 3 to 4 week exposure 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 reeonended
marine criterion of 0.001 ug/1 results.
3/7
-------
REFERENCES CITED;
Borthwick, P.W., et^al., 1973. Accumulation and movement of mi rex
in selected estuaries of South Carolina, 1969-71. Pesticides Monit.
Jour., 7:6.
Bookhout, C.G., e^al., 1972. Effects ofmirex on the larval development
of two crabs. Water, Air, and Soil Pollution, 1:165.
Heath, R. G. and J. W. Spann, 1973. Pesticides in the environment: A
continuing controversy. Intercont. Med. Book Corp., N. Y., pp. 421-435.
Heath, R. G., et^al., 1972. Comparative dietary toxicities of pesticides
to birds. Bureau of Sport Fisheries and Wildlife, Wildlife Report No.
152, U.S. Dept. of the Interior, Washington, D. C., 57 p.
Hyde, K. M., e^ aL , 1974. The effect of mirex on channel catfish
production. Trans. Amer. FishSoc., 103:366.
Kaiser, K.L.E., 1974. Mirex: an unrecognized contaminant of fishes
from Lake Ontario. Science 185:523.
Lowe, J. I., et_ aL , 1971. Effects of mirex on selected estuarine organisms.
In: Transactions of the 36th North American Wildlife Resources Con-
ference, pp. 171-186.
Ludke, J.Li, et^al., 1971. Toxicity of mirex to crayfish, Procambarus
blandingi. Bull. Environ. Contam. Toxicol., 6:89.
Naqvi, S. M. and A. de la Cruz, 1973. Mirex incorporation in the
environment: Residues in non-target organisms, 1972. Pesticides
Monit. Jour., 7:104.
-------
Stickel, W.H., et al., 1973. Pesticides and the environment: A
continuing controversy. Intercont. Med. Book Corp., N. Y., pp. 437-467,
Tagatz, M. E., e£ ah (1975). Seasonal effects of leached mirex on
selected estuarine animals. Arch. Environ. Contam. Toxicol. 3_:371.
Van Valin, C., e£ al., 1968. Some effects of mirex on two warm
water fishes. Trans. Amer. Fish. Soc., 97:185.
Wolfe. J. L. and B. R. Norment, 1973. Accumulation of mirex residues
in selected organisms after an aerial treatment, Mississippi, 1971-72.
Pesticides Monit. Jour., 7:112.
-------
PARATHION
CRITERION:
O.OU ug/1 for freshwater and marine aquatic life.
RATIONALE:
Acute static LC values of the organophosphorus pesticide, parathion,
50
for freshwater fish have ranged generally from about 50 ug/1 for more
sensitive species,such as bluegills, Lepomis macrochirus, to about 2. 5
mg/1 for the more resistant species such as minnows (U.S. Environmental
Protection Agency, 1975). In flowing water exposures, Spacie (1975) obtained
96-hour LC values of 0. 5 mg/1, 1. 6 mg/1, and 1. 76 mg/1 for bluegills,
50
Lepomis macrochirus. fathead minnows, Pimephales promelas, and brook
trout, Salvelinus fontinalis, respectively. Korn and Earnest (1974) found
a 96-hour I£50 of 18 ug/1 for juvenile freshwater and estuarine striped bass
Mororitf saxatilis, in a flowing water system.
Few chronic exposure data are available for aquatic organisms. Brown
bullheads, Ictalurus nebulosus, exposed to 30 ug/1 parathion for 30 days ex-
hibited tremors; at 60 ug/1 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/1, but not at 0.16 ug/1. Tremors,
convulsions, hyper sensitivity, and hemorrhages also were evident at higher
concentrations.
$10
-------
Reproductive impairment and deformities were observed in fathead
minnows exposed to 4. 0 ug/1 for 81/2 months. Development of brook
trout, S.-fontinalis embryos exposed to 32 ug/1 was abnormal and
mortalities associated with premature hatching were observed. Embryos
at 10 ug/1 appeared normal. No adverse effect on juveniles and adults
was evident during 9 months exposure to 7 ug/1.
Inhibition of cholinesterase enzymes is the well established mode of
physiological action of parathion and other organic phosphorous 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 ertal., 1968; Gibson ertal.,
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 ACh E inhibition
ranging from 25 to 90 percent of normal. Weiss (1959) also showed that
susceptibility depended upon the extent of recovery of ACh E 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 exp6sure to 1 mg/1 for 6 to 7 hours (Weiss, 1961).
Some of the other physiological Affects observed to result from exposure
of fish to parathion have been inhibition of spermatogenesis in guppies
Poecilia feticulata, at 10 ug/1 Qttllard and deKinkelin, 1970), alternation of
oxygen consunption rate in bluegills, Leponis nacroGhirusat 100 ug/1 (Dowden,
1966), and liver enlargement associated with increased pesticide-hydrolizing
capability in mosquitofish, Gambusia affinis, Oudke. 1970).
-------
Parathion has been found acutely toxic to aquatic invertebrates at
under one microgram per liter, e. g., a 50-hour LC of 0. 8 ug/1 for
50
Daphnia magna, 48-hour LC of 0. 6 ug/1 for Daphnia pulex, and 48-hour
50
LC of 0. 37 for Simocephalus serrulatus (a daphnid) (Sanders and Cope,
50
1966); a 5-day LC of 0.93 ug/1 for the larval stonefly, Acroneuria pacifica
50
(Jensen and Gaufin, 1964); and a 96-hour LC of 0. 43 ug/1 for the larval
50
caddisfly, Hydropsyche californica (Gaufin et al., 1965). Mulla and
Khasawinah (1969) obtained a 24-hour LC of 0. 5 ug/1 for 4th instar larvae
50
of the midge Tanypus grodhausi. Spacie (1975) obtained 96-hour LCSO's
in flow-through bioassays of 0. 62 ug/1 for Daphnia magna, 0. 40 ug/1 for the
scud, Gammarus fasciatus, and 31. 0 ug/1 for 4th instar of Chironomous
ten tans, a midge. Other invertebrates have been found acutely sensitive
to parathion in concentrations of from 1 to 30 ug/1 in water (U. S.
Environmental Protection Agency, 1975).
Few longer exposures have been conducted. Jensen and Gaiifin (1964)
obtained 30-day LCSO's for Pteronarcys californica and Acroneuria pacifica
of 2. 2 and 0. 44 ug/1, respectively. Spacie (1975) found the three-week LC
50
for Daphnia magna to be 0.14 ug/1. Statistically significant reproductive
impairment occurred at concentrations above 0. 08 ug/1. A 43-day LC
50
of 0. 07 ug/1 was reported for Gammarus fasciatus and a concentration of
0.04 ug/1 produced significantly greater mortality than among controls.
Limited information is available on persistence of parathion in water.
Eichelberger and Lichtenberg (1971) determined theffyalf-life in river water
(pH 7. 3 - 8. 0) to be one week. Using AChE inhibitory capacity as the
indicator, Weiss and Gakstatter (1964) found the half-life of parathion
32
-------
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, Gambusia 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, Palaemonetes kadiakensis,
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, Ictalurus-melas, 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, Lepomis macrochirus, had about 25 times
water concentrations in their bodies. Leland (1968) demonstrated a biological
half-life of parathion in rainbow trout, Salmo gairdneri, 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.
Weiss and Gakstatter (1964) have shown that 15-day continuous
exposure to parathion (1. 0 ug/1) 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 (Eaton, 1970) and for short
periods (Carter, 1971; Coppage and Duke, 1971; Coppage, 1972; Coppage
and Matthews, 1974; Post and Leasure, 1974; Coppage et^al. , 1975). It
has been shown that a concentration of 10 ug parathion/1 of flowing
sea water kills 40 to 60 percent of the marine fishes Lagodon rhomboides
(pinfish) and Leostomus xanthurus (spot) in 24 hours and causes about
87 to 92 percent brain AChE inhibition (Coppage and Matthews, 1974).
Similar inhibition of ACJCE and mortality were caused in sheepshead minnows,
Cyprinodon variegatus, in 2, 24, 48, and 72 hours at concentrations of 5000,
2000, 100, and 10 ug/1, respectively in static tests (Coppage, 1972). These
data indicate that reductions of brain ACllE activity of marine fishes by
70 to 80 percent or more in short-term exposures to parathion may be
associated with some deaths.
Other estimates of parathion toxicity to marine organisms follow. The
48 -hour EC for parathion to Penaeus duorarum was found to be 0. 2 ug/1
50
(Lowe et al. , 1970). Lahav and Sarig (1969) reported the 96 -hour LC
50
for mullet, Mugil cephalus, to be 125 ug/1. 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/1.
An application factor of O.I is applied to the 96-hour LC50 data
for invertebrates which range upward from O.k ug/1. A criteria of O.
is recommended for marine and freshwater aquatic life.
-------
REFERENCES CITED;
Billard, R. and deKinkelin, 1970. Sterilization of the testicles of
guppies by means of non-lethal doses of parathion. Annales D
Hydrobiologie, 1(1):91.
Burke, W.D. and D. E. Ferguson, 1969. Toxicities of four insecticides
to resistant and susceptible mosquitofish in static and flowing solutions.
Mosquito News, 29(1):96.
Carter, F. L., 1971. In vivo studies of brain acetylcholinesterase
inhibition by organophosphate and carbamate insecticides in fish.
Ph. D. dissertation, Louisiana State University, Baton Rouge, La.
Coppage, D. L., 1972. Organophosphate pesticides: specific level of
brain ACHE inhibition related to death in sheepshead minnows.
Trans. Amer. Fish. Soc., 101:534.
Coppage, D. L. and T. W. Duke, 1971. Effects of pesticides in estuaries
along the Gulf and Southeast Atlantic Coasts. In_: Proceedings of the
2nd Gulf Coast Conference on Mosquito Suppression and Wildlife
Management (C. H. Schmidt, ed.) National Mosquito Control - Fish
and Wildlife Management Coordinating Committee, Washington, D. C.
Coppage, D. L. and E. Matthews, 1974. Short-term effects of
organophosphate pesticides on cholinesterases of estuarine fishes
and pink shrimp. Bull. Environ. Contain. Toicicol., 11:483.
-------
Coppage, D. L. et al., 1975. Brain acetylcholinesterase inhibition
in fish as a diagnosis of environmental poisoning by malathion,
O, O-dimethyl S-(l, 1-dicarbethoxy-ethyl) phosphorodithioate.
Pesticide Biochemistry and Physiology (in press).
Dowden, B. F., 1966. Effects of five insecticides on the oxygen
consumption of the bluegill sunfish, Lepomis macrochirus.
Ph. D. Thesis, Louisiana State University, Baton Rouge, La.
Eaton, J. G., 1970. Chronic malathion toxicity to the bluegill, '
Lepomis macrochirus, Rafinesque). Water Research. 4:673.
Eichelberger, J. W. and J. J. Lichtenberg, 1971. Persistence
of pesticides in river water. Environ. Sci. and Technol., 5:541.
Gaufin, A. R., et^aL, 1965. Toxicity of ten organic insecticides to
various aquatic invertebrates. Water and Sew. Works, 112:276.
Gibson, J. R., ertal., 1969. Sources of error in the use of fish-brain
acetylcholinesterase activity as a monitor for pollution. Bull.
Environ. Contam. Toxicol., 4:17.
Jensen, L. D. and A. R. Gaufin, 1964. Long-term effects of organic
insecticides on two species of stonefly naiads. Trans. Amer. Fish.
Soc., 93:357.
Korn, S. and R. Earnest, 1974. Acute toxicity of twenty insecticides to
striped bass, Morone saxatilis. Calif. Fish and Game., 60:128.
-------
Lahav, M. and S. Sarig, 1969. Sensitivity of pond fish to cotnion
(azinphosmethyl) and parathion. Bamidgeh, Bull. Fish. Cult.
Israel, 21:67.
Leland, H. V. II, 1968. Biochemical factors affecting toxicity of
parathion and selected analogs to fishes. University of Michigan,
Ann Arbor, Michigan.
Lowe, J. I., et^aL,, 1970. Laboratory bioassays. In_: Progress
Report for Fiscal Year 1969, Pesticide Field Station, Gulf Breeze,
Florida. U. S. Fish. Wildl. Serv. Circ. 335.
i
Ludke, J. L., 1970. Mechanism of resistance to parathion in mosquitofish,
Gambusia affinis. Ph.D. Thesis, Mississippi State University,
University, Mississippi.
Mount, D. I. and H. W. Boyle, 1969. Parathion - Use of blood
concentration to diagnose mortality of fish. Environ. Sci. and
Technol., 3:1183.
Mulla, M. S. and A. M. Khasawinah, 1969. Laboratory and field
evaluation of larvicides against Chironomic midges. Jour. Econ.
Entomol., 62:37.
Murphy, S. D., e£al_., 196.8. Comparative anticholinesterase action •
of organophosphorus insecticides in invertebrates. Toxicol.
i
Appl. pharmacol., 12:22.
Post, G. and R. A. Leasure, 1974. Sublethal effect of malathion to
three salmonid species. Bull. Environ. Contain. Toxicol., 12:312.
-------
Sanders, H. O., 1972. Toxicity of some inecticides to four species of
malacostracan crustaceans. U. S. Department of the Interior,
Washington, D. C., Bureau of Sport Fisheries and Wildlife,
Technical Paper No. 66.
Sanders, H. O. and O. B. Cope, 1966. Toxicities of several pesticides
to two species of Cladocerans. Trans. Amer. Fish. Soc.., 95:165.
Spacie, A., 1975. Acute and chronic parathion toxicity to fish and
invertebrates. U. S. Environmental Protection Agency, EPA
Ecological Research Series. (In preparation).
U. S. Environmental Protection Agency, 1975. Initial scientific and
minieconomic review of parathion. U. S. Environmental Protection
Agency, Office of Pesticides Programs, Report. No. EPA-540/1-75-001.
National Technical Information Service, Springfield, Virginia.
Weiss, C. M., 1958. The determination of cholinesterase in the brain
tissue of three species of freshwater fish and its inactivation in vivo.
Ecology, 39:194.
Weiss, C. M., 1959. Response of fish to sublethal exposures of organic
phosphorus insecticides. Sew. and Ind. Wastes, 31:580.
Weiss, C. M., 1961. Physiological effect of organic phosphorus
insecticides on several species of fish. Trans. Amer. Fish. Soc., 90:143;
' i
f
Weiss, C. M. and J. H. Gakstatter, 1964. The decay of anticholinesterase
activity of organic phosphorus insecticides on storage in waters of
different pH. Advances in Water Poll. Res., 1:83.
-------
TOXAPHENE
CRITERIA;
5 ug/1 for domestic water supply (health);
0.005 ug/1 for freshwater and marine aquatic life.
RATIONALE;
The highest level of toxaphene found to have minimal or no
long-term effects in the most sensitive mammal tested, the dog,
is 10.0 mg/kg in the diet or 1.7 mg/kg of body weight/day
(Lehman, 1965). Where adequate human data are not available for
corroboration of the animal results, the total "safe"
intake level is assumed to be 1/500 of the "no effect" or
"minimal effect" level reported for the most sensitive animal
tested.
Applying the available data and based upon the assumption
that 20 percent of the total intake of toxaphene is from drinking
water, that the average person weighs 70 kg and consumes 2 liters
of water per day, the formula for calculating a criterion is 1.7
mg/kg X 0.2 X 70 kg X 1/500 X 1/2 =0.024 mg/1. However, at 0.024 mg/1
there is an organoleptic effect which has been shown to occur at the level of 0.005
mg/1 (Cohen, et-al., 1961). Thus the criterion is set at 5 ug/1.
Macek and McAllister (1970) reported 96-hour LC50 values for 12
fish species ranging from 2 ug/1 for largemouth bass, Micropterus
-------
salmoides, to 18 ug/1 for bluegill, Lepomis macrochirus. Mahdi
(1966) reported 96-hour LC50 values as low as 1.8 ug/1 for black
bullhead, Ictalurus melas, and as high as 50 ug/1 for goldfish,
Carrassius auratus, and Henderson, et al. (1959) reported 96-hour
TLra values from 3.5 ug/1 for bluegill to 20 ug/1 for guppies,
Poecilia reticulata. The 96-hour LC5Q values for three stonefly
species, Pteronarcys californica, Pteronarcella badia, and
Claassenia sabulosa, were reported by Sanders and Cope (1968) to
range from 1.3 to 3.0 ug/1.
In a chronic continuous flow bioassay with brook trout,
Salvelinus fontinalis, Mayer, et al. (1975) found that-toxaphene
in water at a level of 0.039 ug/1 adversely affected the growth
and development of brook trout fry. The mortality of fish
at this level was significantly greater than the controls,
indicating a no-effect level of less than 0.039 ug/L The
96-hour LC of toxaphene for 16-month-old trout was 10.8
50
ug/1. Mehrle and Mayer (1975) conducted a series of long-term
studies on the fathead minnow , Pimephales promelas. exposing the
organisms for 150 days to concentrations of toxaphene as low
as 0.055 ug/1 in a flow-through system . Their results
confirmed the Mayer, e_t aX ,(1975) work, showing that growth of
all fish exposed to concentrations of 0.055 ug/1 was
significantly reduced. A no-effect level for this freshwater
fish would be less than 0.055 ug/1.
-------
Hughes (1968) reported that lakes treated with toxaphene
concentrations ranging from 40 to 150 ug/1 remained toxic to fish
for periods of a few months to five years., Terriere, et al.
(1966) reported that a lake treated with toxaphene as a piscicide
remained toxic to fish for at least five years. Bioconcentration
factors of toxaphene were 500 for aquatic plants, 1,000 to 2,000
for aquatic animals other than fish, and 10,000 to 20,000 for
rainbow trout in the lake. flayer, at al. (1975) observed
accumulations of 5,000 to 21,000 times water concentrations in
brook trout exposed only through the water. Accumulation factors
of 3,400 to 17,000 from aqueous solution have been reported for
bacteria, algae and fungi (Paris, et al. , 1975).
Heath, et al. (1972) reported the 5-day dietary LC5 to be 96
to 142 ppm for young birds of four species. IIo reproductive
impairment occurred in mallards resulting from a long-term
dietary dosage of 7 ppm and starlings tolerated 45 ppm for long
periods (Patuxent Wildlife Research Center, Laurel, Maryland,
unpublished data). At water levels known to affect fish (0.04
ug/1, Mayer, et aL, 1975)and with accumulation -factors similar to those cited above.
(20,000 times), resulting residue levels (0.04 x 20,000- 0.8 ppm) would not be
expected to approach dosage levels known to be a hazard to birds.
No guideline for toxaphene has been set by the U.S. Pood and
Drug Administration as a residue limit for edible tissues of fish
for human consumption.
-------
Lowe (19C4) reported a 24-hour LC5Q of 3.2 ug/1, a 48-hour LC5o
of 1.0 ug/1, and a 144-hour LCsg of 0.5 ug/1 for the spot,
Leiostoraus xanthurus . Butler (1963) reported a 90.8 percent
decrease in productivity of natural phytoplankton communities
during a 4-hour exposure to a concentration of 1,000 ug/1 of
toxaphene. Lowe also reported a 48-hour EC of 4.9 ug/1 for the
brown shrimp, Penaeus aztecus; and a 48-hour ECso of 330 ug/1 for
juvenile blue crabs, Callinectes sapidus. A concentration of 57
ug/1 resulted in a 50 percent decrease in oyster, Crassest re a
virginica, shell grov/th after 96 hours of exposure at a water
temperature of 31° C and 24 o/oo salinity, whereas 63 ug/1
produced the same effect at 19° C and 19 o/oo salinity. The
48-hour LC5Q for the juvenile white mullet, Mugil cure ma, was 5.5
ug/1.
Korn and Earnest (1974)^ rejx>rtc a, 9^-hoiy: LC^ for the
striped bass, Morone saxatilis of 4.4 ug/iwlhe 96-hour LC for the
50
pinfish, Lagodon rhomboides, an organism of wide geographic distribution
and ecological inportance(Caldwell, 1957) , has been reported as 0.5 ug/1
(Schiimel, et al. , in preparation) . While the u»e of an application
factor of 0.01 has been recommended by the NAS-NAE (NAS, 1974) its
use is especially appropriate in the case of toxaphene because long-
term studies with fathead minnows, Pimephales promelas (Mehrle and
Mayer, 1975) , and brook trout, Salvelinus fontinalis (Mayer', et al. , 1975) ,
have failed to establish a no-effect level. Application of the 0.01
factor to the 96-hour LC for the pinfish yields a marine criterion of
0.005 ug/1.
-------
A no-effect level was not achieved in the studies using
freshwater organisms. Mortality and adverse physiological and
physical effects were detected at the lowest concentration used,
0.039 ug/1. Hence, for toxaphene, the use of the same criterion
for botn marine and freshwater is recomvended.
33.3
-------
REFERENCES CITED:
Butler, P.A., 1963. Pesticide-wildlife studies. A review of
Pish and Wildlife Service Investigations during 1961 and
1962. U.S. Dept. of the Interior, Fish and Wildl. Ser.,
Circ. 167.
Caldwell, D.K. 1957. The biology and systematics of the
plnfish » Lagodon rhomboides (Dinmaeus). Bull. Florida
State Mus. 2:1.
Cohen, J.M., et al., 1961. Effects of fish poisons on water
supplies. Jour. Amer. Water Works Ass., 53:49.
Heath, R.G., et jO_., 1972. Comparative dietary toxicity of
pesticides to birds. U.S. Dept. of the Interior, Bur.
of Sport Fish, and Wildl., Washington, D.C., Wildlife
Rept. No. 152.
Henderson, D., et ajl., 1959. Relative toxicity of ten
chlorinated hydrocarbon insecticides to four species of
fish. Trans. Amer. Fish. Soc., 88:23.
Hughes, R.A., 1968. Persistence of toxaphene in natural
waters, M.S. Thesis, University of Washington, Seattle.
Korn, S. and R. Earnest, 1974. Acute toxicity of twenty
insecticides to striped bass, Morone a_axatili8_. Calif.
Fish and Game, 60:128.
-------
Lehnan, A.J., 1965. Summaries of pesticide toxicity. Assn.
of Food and Drug Officials of «-he U.S., Topeka, Kansas,
pp. 1-40.
Loiref J.I., 1964. Chronic exposure of the spot, Leiostomus
xanthurus, to sublethal concentrations of toxaphene in
seawater, Tran. Amer. Fish Soc. 93:396.
Macek, K.J. and W.A. McAllister, 1970. Insecticide
susceptibility of some common fish family
representatives. Trans. Amer. Fish. Soc., 1:20.
Mahdi, II.A., 1-966. Mortality of some species of fish to
toxaphene at three temperatures. U.S. Dept. of the
Interior. Bur. of Sport Fish, and Wildl., Investigation
in Fish Control, 6:1.
Mayer, F.L., Jr., et al.,. Toxaphene effects on
A
reproduction, growth and mortality of brook trout. Fish
Pesticide Research Laboratory, Columbia, Mo., U.S.
Environmental Protection Agency Contract No.
EPA-IAG-0153.
Mehrle, P.M. and F.L. Mayer, 1975. Toxaphene effects on growth
and bone composition of fathead minnow Pimephales promel*a
Jour. Fish. Res. Bd. Canada. 32:593.
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972. U.S. Government Printing Office,
Washington, D.C.
-------
Paris, D.F., et_ al_. , 1975. Microbial degradation and accumulation
of pesticides in aquatic systems. U.S. Environmental
Protection Agency, Washington D.C., EPA-660/3-75-007 .
Patuxent wildlife Research Center. 1974. Toxaphene research
data. U.S. Dept of Interior, Bur. of Sport Fish, and
Wildlife, (Unpublished).
Sanders, H.O. and O.B. Cope, 1968. The relative toxicities
of several pesticides to naiads of three species of
stoneflies. Lirunol. Oceanog. , 13:112.
, S.C.. &*
^ . No. 2.69.
Environmental Research Laboratory, contribution
LC etal., 1966. The persistence of toxaphene in
Terriere, L.C., e_c »JL. , ar,
-------
CRITERIA;
Range
5 - 9 Domestic water supplies (welfare);
6. 5 - 9. 0 Freshwater aquatic life;
6. 5 - 8. 5* Marine aquatic life.
INTRODUCTION;
"pH" is a measure of the hydrogen 1on activity in a water
sample. It is mathematically related to hydrogen ion activity accord-
+ +
ing 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 (CO ), carbonic acid (H CO ),
2 _ 23
bicarbonate ion (HCO ), and carbonate ions (CO** ). The interactions and
3 3
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 fey changes in pH. This effect is important because the toxicity
of many compounds is affected by the degree of dissociation. One such
*. .. .but not more than 0. 2 units outside of normally occurring range.
-------
example is hydrogen cyanide (HCN). Cyanide toxicity to fish increases
as the pH is lowered because the chemical equilibrium is shifted
toward an increased concentration of HCN. Similar results have been
shown for hydrogen sulfide (H S) (Jones, 1964).
2
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 equili-
brium between hypochlorous acid (HOC1) and the hypochlorite ion
(OC1~) according to the reaction:
HOC1 = H+ + OC1-
Butterfield (1948) has shown that chlorine disinfection is more effect-
ive 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 maxi-
mum 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 freshwater fish has been published
by the European Inland Fisheries Advisory Commission (EIFAC, 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 CO from bicarbonate in the
water either to be directly toxic, or to cause the pH range 5 - 6 to
become lethal. "
Mount (1973) performed bioassays on the fathead minnow, 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 100 ppm.
6. 5 - 9. 0 Harmless to fish, although the toxicity of other poisons
may be affected by changes within this range.
* EIFAC, 1969.
-------
5. 9, 6. 6, and a control of 7. 5. At the two lowest pH values (4. 5
and 5. 2) behavior was abnormal and the fish .were deformed. At pH
values less than 6. 6, egg production and egg 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), and stoneflies (four species), dragonflies (two species), and
mayflies (one species). All are important fish food organisms. The
30-day TLgg values ranged from 2. 45 to 5. 38 with the caddisflies
being the most tolerant and the mayflies 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 invertebrate 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 concen-
trations which are also toxic. Ammonia has been shown to be ten
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 sea water, the naturally occurring variability of pH is less
than in fresh water. Some marine communities are more sensitive
to pH change than others (NAS, 1974). Normal pH values in sea water
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
3v^
-------
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 pH 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 the
freshwater criteria, rapid pH fluctuations that are due to 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 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 approxi-
mately 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.
-------
REFERENCES CITED:
Bell, H.L., 1971. Effect of low pH on the survival and emergence
of aquatic insects. Water Res., 5:313.
Butterfield, C.T., 1948. Bacterial properties of free and combined
available chlorine. Jour. Amer. Water Works Assn., 40:1305.
Capurro, L.R.A., 1970. Oceanography for Practicing Engineers, Barnes
and Noble, Inc., New York.
Delfino, J.J. and G.F. Lee, 1971. Variation of manganese, dissolved
oxygen and related chemical parameters in the bottom waters of Lake
Mendota, Wisconsin. Water Res., 5:1207.
European Inland Fisheries Advisory Commission, 1969. Water quality
criteria for European freshwater fish--extreme pH values and inland
fisheries, prepared by EIFAC Working Party on Water Quality Criteria
for European Freshwater Fish. Water Research, 3:593.
Jones, J.R.E., 1964. Fish and River Pollution. Butterworth, London.
Langelier, W.F., 1936. The analytical control of anti-corrosion
water treatment. Jour. Amer. Water Works Assn., 28:1500.
Mount, D.I., 1973. Chronic effect of low pH on fathead minnow survival,
growth and reproduction. Water Res., 7:987.
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972. U.S. Government Printing Office,
Washington, D. C.
-------
Reid, L.C. and D.A. Carlson, 1974. Chlorine disinfection of low
temperature waters. Jour. Enviro. Engineering Div., ASCE,
100: No. EE2, Proc.-Papers 10443, 339.
Sawyer, C.N., 1960. Chemistry for sanitary engineers. McGraw-Hill,
New York.
Stumm, W. and J.J. Morgan, 1970. Aquatic Chemistry. John Wiley
and Sons, Inc., New York, Chapter 4.
Zirino, A. and S. Yamamoto, 1972. A pH-dependent model for the chemical
specialization of copper, zinc, cadmium and lead in seawater.
Limn, and Oceanog., 17(5).
3V*
-------
PHENOL
CRITERION:
1 ug/1 for domestic water supply (welfare), and
to protect against fish flesh tainting.
INTRODUCTION:
Phenolic compounds include a wide variety of organic chemicals.
The phenols may be classified into monohydric, dihydric, and
poiyhydric phenols depending upon the number of hydroxyl groups
attacned to the aromatic ring. Phenol itself, which has but one
hydroxy'l group, is the most typical of the group and is often
»sed as a model compound. The properties of phenol, with certain modifi-
cations depending on the nature of the substituents on the benzene ring, are
shared by other phenolic compounds. Phenolic compounds arisefrom the distill-
ation of coal and wood; from oil refineries; chemical plants;.livestock
dips; human and other organic wastes; hydrolysis, chemical oxidation
and microbial degradation of pesticides; and from naturally occurring
sources and substances. Some compounds are refractory to biological
degradation and can be transported long distances in water.
RATIONALE:
Phenolic compounds can affect freshwater fishes adversely by
direct toxicity to fish and fish-food organisms; by lowering the amount
of available oxygen because of the high oxygen demand of the compounds
and by tainting of fish flesh (EIFAC, 1973). Shelford (1917) observed
-------
that a concentration of 1 cc per liter (purity of compound and concen-
tration are unknown) was rapidly fatal to fish but solutions of one half
to three quarters of this amount (i.e., .5 to .75 cc) would require up to
one hour to kill fish. Subsequent studies have confirmed the toxicity
of phenol to both adult and immature organisms (EIFAC, 1973). Decreased
egg development in the oyster, Crassostrea virginica, has been found to
occur at levels of 2 mg/1 phenol (Davis and Hidu, 1969).
Various environmental Conditions will increase the toxicity of phenol.
Lower dissolved oxygen concentrations, increased salinity and Increased
temperature all enhance the toxicity of phenol (EIFAC, 1973). It has been
shown that phenol and o-cresol have 24-hour LCBO's of 5 and 2 mg/1
respectively for trout embryos (Albersmeyer and von Erichsen, 1959).
Rainbow trout were killed in 7.3 mg/1 phenol in 2 hours and in 6.5 mg/1
phenol in 12 hours; at these concentrations there was rapid damage to
x.
gills and severe pathology of other tissues (Mitrovic, ejt al_., 1968).
Pathologic changes in gills and in fish tissues were found at concentrations
-------
2 yg/1 for 2-chlorophenol (Burttschell, e_t al_., 1959). The chlorinated
phenols present problems in drinking water supplies because phenol is
not removed efficiently by conventional water treatment and can be
chlorinated during the final water treatment process to form persistent
odor-producing compounds. Thus, odor problems are created in the
distribution system. Boetius (1954), Fetterolf (1962), Schulze (1961),
and Shumway (1966) estimated threshold fish flesh tainting concentrations
for o-chlorophenol, p-chlorophenol, and 2, 4-dichlorophenol to range
from 0.1 ug/1 to 15 ug/1. The O-chlorophenol produced tainting at the
lower concentration.
A criterion of 1 ug/1 phenol, which is about half of the chlorophenol
odor effect level for a water supply and near the threshold fish flesh
tainting concentration, should protect the freshwater environment for
such users.
-------
REFERENCES CITED;
Albersmeyer, W. and L. von Erichsen, 1959. Untersuchungen zur wlrung
von Teerbestandteilen in abwassern. MHteilungen, Z. F1sch., 3:29.
Boetius, J., 1954. Foul taste of fish and oysters caused by chlorophenol.
Medd. Dan. Fisk. Havunders. 1:1.
Burttschell, et a]_., 1959. Chlorine derivatives of phenol causing taste
and odor. Jour. Amer. Water Works Assn., 51:205.
Davis, H.C. and H. Hidu, 1969. Effects of pesticides on embryonic development
of clams and oysters and on survival and growth of the larvae. Fish.
Bull., 67:393.
European Inland Fisheries Advisory Commission, 1973. Water quality criteria
for European freshwater fish. Report on mono-hydric phenols and inland
fisheries. Water Res., 7:929.
Fetterolf, C.M., 1962. Investigation of fish off flavor, Muskegon
lake. Bur. of Water Management, Michigan Department of Natural
Resources, Lansing, Michigan.
McKee, J.E. and H.W. Wolf, 1963. Water quality criteria. State Water
Quality Control Board, Sacramento, California. Pub. 3-A.
Mitrovic, V.A., et aJL, 1968. Some pathologic effects of sub-acute and
acute poisoning of rainbow trout by phenol in hard water. Water
Res., 2:249.
Reichenbach-Klinke, H.H., 1965. Der phenolgeholt des wasser in seiner
auswerkung auf den fischorgaismus. Arch. Fischwiss., 16:1.
35-*
-------
tosen, A.A., et al.., 1962. Odor thresholds of mixed organic chemicals,
Jour. Water Poll. Cont. Fed., 34:7.
Schulze, E., 1961. The Influence of phenol-containing effluents on the
taste of fish. Int. Rev. ges. Hydroblol. 46:84.
She!ford, V.E., 1917. An experimental study of the effects of gas wastes
upon fishes, with especial reference to stream pollution. Bull, of
the Illinois State Lab. of Nat. Hist., XI:381.
Shumway, D.I., 1966. Effects of effluents on flavor of salmon flesh.
Agricultural Experiment Station, Oregon State University,
CorvalHs, 17 p.
-------
PHOSPHORUS
CRITERION:
0.10 ug/1 yellow (elemental) phosphorus for marine or
estuarine waters.
INTRODUCTION
Phosphorus 1n the elemental form 1s particularly toxic and 1s subject to
bioaccumulation In much the same way as mercury. Phosphorus as phosphate 1s
one of the major nutrients required for plant nutrition and 1s 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, 1t 1s recognized that phosphorus
1s not the sole cause of eutrophication but there 1s substantiating evidence
that frequently 1t 1s the key element of all of the elements required by
freshwater plants, and generally, 1t 1s present 1n the least amount relative
to need. Therefore? an Increase 1n phosphorus allows use of other already
present nutrients for plant growth. Further, of all of the elements required
for plant growth 1n the water environment, phosphorus 1s the most easily con-
trolled by man.
Large deposits of phosphate rock are found near the western shore of
Central Florida, as well as 1n a number of other States. Deposits 1n Florida
are found in the form of pebbles which vary 1n size from fine sand to about
the size of a human foot. These pebbles are embedded in a matrix of clay and
352.
-------
and sand. The phosphate rock beds He within a few feet of the surface and
"rifling 1s accomplished by use of hydraulic water jets and a washing operation
that separates the phosphate from waste materials. The process 1s similar to
that of strip-mining. Florida, Idaho, Montana, North Carolina, South Carolina,
Tennessee, Utah, Virginia, and Wyoming share phosphate mining activities.
Phosphates enter waterways from several different sources. The human
body excretes about one pound per year of phosphorus expressed as "P." The
use of phosphate detergents and other domestic phosphates Increases the per
capita contribution to about 3-1/2 pounds per year of phosphorus as P. Some
Industries, such as potato processing, have wastewaters high 1n phosphates.
Crop, forest, Idle, and urban land contribute varying amounts of phosphorus
diffused sources 1n drainage to watercourses. This drainage may be surface runoff
ef rainfall, effluent from tile lines, or return flow from Irrigation. Cattle
feedlots, concentrations of domestic duck or wild duck populations, tree leaves,
and fallout from the atmosphere all are contributing sources.
Evidence Indicates that: (1) high phosphorus concentrations are associated
with accelerated eutrophlcation of waters, when other growth-promoting factors
are present; (2) aquatic plant problems develop 1nreservoirs and other standing
waters at phosphorus values lower than those critical 1n 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 reservoir; (2)
the volume of the euphotic zone; (3) the extent of biological activities;
3*3
-------
(4) the detention time within the lake basin or the time.avail able for biological
activities; and, (3) the level of discharge from the lake or of the penstock
from the reservoir.
Once nutrients are combined within the aquatic ecosystem, their removal 1s
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 blotic community, while the remainder may be
deposited with sediments. Much of the material that becomes combined 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 LC50 of 0.105 mg/1 at 48 hours and 0.025 mg/1 at
160 hours for blueglll sunfish, Lepomis macrochirus, exposed to yellow phos-
phorus in distilled water at 26° C and pH 7. The 125- and 195-hour LCSO's
of yellow phosphorus to Atlantic cod, Gadus morhua, and Atlantic salmon,
Salmo salar smolts 1n continuous-exposure experiments was 1.89 and 0.79 ug/1,
respectively.(Fletcher and Hoyle, 1972). No evidence of an incipient lethal
level was observed since the lowest concentratiori of P4 tested was 0.79 ug/1.
Salmon that were exposed to elemental phosphorus concentrations of 40 ug/1 or
less developed a distinct external red color and showed signs of extensive
hemolysis. The predominant features of ?4 poisoning in salmon were external
redness, hemolysis, and reduced hematocrlts.
3S1/
-------
Following the opening of an elemental phosphorus production plant in
Long Harbour, Placentia Bay, Newfound!arid, divers observed dead fish upon the
bottom throughout the Harbour (Peer, 1972). Mortalities were confined to a
water depth of less than 18 meters. There was visual evidence of selective
mortality among benthos. Live mussels were found within 300 meters of the
effluent pipe, while all scallops within this area were dead.
Fish will concentrate elemental phosphorus from water containing as
little as 1 ug/1 (Idler, 1969). In one set of experiments, a cod swimming in
water containing 1 ug/1 elemental phosphorus for 18 hours concentrated phos-
phorus to 50 ug/kg in muscle, 150 ug/kg in fatty tissue, and 25,000 ugAg in
the liver (Idler, 1969 ; Jangaard, 1970). The experimental findings showed
that phosphorus is quite stable in the fish tissues.
The criterion of 0.10 ug/1 elemental phosphorus for marine or estuarine
waters is 1/10 of demonstrated lethal levels to important marine organisms
and of levels that have been found to result in significant bioaccumulation.
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/1 P may
interfere with coagulation in water treatment plants. When such concentrations
exceed 25 ug/1 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 impart 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 ac-
celerated or cultural eutrophication, total phosphates as phosphorus (P)
should not exceed 50 ug/1 in any stream at the point where it enters any lake
or reservoir, nor 25 ug/1 within the lake or reservoir. A desired goal for
the prevention of plant nuisances in streams or other flowing waters not dis-
charging directly to lakes or impoundments is 100 ug/1 total P (Mackenthun, 1973),
Most relatively uncontaminated lake districts are known to have surface waters
that contain from 10 to 30 ug/1 total phosphorus as P (Hutchinfon, 1957).
The majority of the Nation's eutrophication problems are associated
with lakes or reservoirs and currently there are more data to support
the establishment of a limiting phosphorus level in those waters than in
streams or rivers that do not directly impact such water. There are
natural conditions, also, that would dictate the consideration of either
-------
a more or less stringent phosphorus level. Eutrophication problems may
occur in waters where the phosphorus concentration is less than that
indicated above and, obviously, there would be a need in such waters to
have nutrient limits that are more stringent. 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 1n 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 down-
stream, 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.
357
-------
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 *n meters is divided by the hydraulic detention time
in years. Vollenweider's data suggest a range of leading values that should
result in oligotrophic lake water quality.
Mean Depth/Hydraulic
Retention Time
Oligotrophic or
Permissible
Loading
Eutrophic
or Critical
Loading
(meters/year) (grams/meter2/year) (grams/meter2/year)
0.5
1.0
2.5
5.0
7.5
10.0
25.0
50.0
75.0
100.0
0.07
0.10
0.16
0.22
0.27
0.32
0.50
0.71
0.87
1.00
0.14
0.20
0.32
0.45
0.55
0.63
1.00
1.41
1.73
2.00
There may be waterways wherein higher concentrations or loadings of total
phosphorus do not produce eutrophy, as well as those waterways wherein lower
concentrations or loadings of total phosphorus may be associated with popula-
358
-------
tlons 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.
Often, naturally occurring phenomena limit the development of plant nuisances;
often there are technological or cost-effective limitations to the control of
introduced pollutants. Exceptions to the threat of phosphorus in eutrophication
occur in waters highly laden ' with natural silts or colors which reduce the
penetration of sunlight needed for plant photosynthesis; in those waters whose
morphometric features of steep banks, great depth, and substantial flows
contribute to a history of no plant problems; in those waters that are managed
primarily for waterfowl or other wildlife; 1n those waters where an identified
nutrient other than phosphorus is limiting to plant growth and the level and
nature of such limiting nutrient would not be expected to increase to an extent
that would influence eutrophication; and in those waters where 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.
3s?
-------
REFERENCES CITED
Fletcher, G.L., and R.O. Hoyle, 1972. Acute toxlcity of yellow phosphorus
to Atlantic cod (Gadus morhua) and Atlantic salmon (Salmo salar) smolts.
Jour. Fish. Res. Bd. of Canada, 29:1295.
Hutchinson, G.E., 1957. A treatise on limnology. John Wiley & Sons, New York.
Idler, D.R., 1969. Coexistence of a fishery and a major industry in Placentia
Bay. Chemistry in Canada, 21(11):16.
Isom, B.G., 1960. Toxicity of elementary phosphorus. Jour. Water Poll.
Control Fed., 32:1312.
Jangaard, P.M., 1970. The role played by the Fisheries Research Board of
Canada in the "red" herring phosphorus pollution crisis in P'lacentia Bay,
Newfoundland. Fisheries Research Board, Atlantic Regional Office (Circular
No. 1) Halifax, Nova Scotia.
Mackenthun, K.M., 1973. Toward a cleaner aquatic environment. Environmental
Protection Agency, Washington, D.C.
Peer, D.L., 1972. Observations on mortalities of benthic organisms after
contamination of the bottom of Long Harbbur, Placentia Bay, Newfoundland
with elemental phosphorus. In: Effects of Elemental Phosphorus on Marine
Life, Fish. Res. Bd. of Canada, Circular 2, pp. 181-186.
Vollenweider, R.A., 1973. Input output models. Schweiz Z. Hydro!. (In Press)
-------
PHTHALATE ESTERS
CRITERION;
3 ug/1 for freshwater aquatic life.
INTRODUCTION:
Phthalate esters are organic compounds extensively used as plastidzers,
particularly in polyvlnyl chloride plastics. The function of a plasticlzer
1s to change the characteristics of the plastic resin by making them
more flexible or to Improve their workability. Some plastic formulations may
contain up to 60 parts per hundred of a phthalate ester. Several" phthalate
esters are synthesized and vary 1n the side chain length and structure attached
to the parent benzene ring. Certain esters, the d1-2-ethylhexyl and di-n-butyl
phthalates, are used as an orchard acaricide and Insect repellent, respectively.
Occurrence of the phthalate residues has been demonstrated in fresh
(Mayer, e_t a_l_., 1971), marine (Morris, 1970), industrial and heavily populated
waters (Stalling, 1972). No well documented information exists on the fate of
the phthalate compounds 1n aquatic environments (Mayer, e_t al_., 1972).
RATIONALE:
Acute toxicity of the phthalate esters tested has been shown to be quite
low (Sanders ejt a_K, 1973). Mayer and Sanders (1973) determined the 96-hour LC50
of di-n-butyl phthalate to be 1.3 mg/1 for the fathead minnow, Pimephales
promelas; 0.73 mg/1 for the bluegill, Lepomis macrochirus; 2.91 mg/1 for channel
catfish, Ictalurus punctatus; 6.47 mg/1 for rainbow trout, Salmo gairdnerl;
2.10 mg/1 for the scud, Gammarus pseudolimnaeus; and >10 mg/1 for the crayfish,
Orconectes nais. Daphm'a magna, when exposed to the di-2-ethylhexyl phthalate at
-------
levels of from 3 ug/1 to 30 ug/1 for 21 days, exhibited a decrease in the nutnbe
of young produced of 60 to 83 percent, respectively.
Ability of aquatic organisms to accumulate various phthalate residues
depend upon the ester, concentration and time of exposure. Concentration
factors up to 6,600 have been reported for di-n-butyl phthalate in midge
larva, Cnironomus plumosus, after seven days exposure to 0.18 ug/1. Amphipods,
Gammarus pseudolimnaeus, were found to concentrate the di-2-ethylhexyl
phthalate 13,400 times in water containing 0.1 ug/1 after 14 days exposure
(Sanders e£al_., 1973). Daphnia magna were exposed to 0.1 ug/1 of radioactively
labeled phthalate for 7 days and then transferred to fresh flowing water to determine
the time required for elimination of phthalate residues (Mayer and Sanders, 1973).
After 3 days, 50 percent of the total radioactivity remained; 25 percent of the
activity was still present after 7 days in fresh water.
Phthalate esters can be detrimental to aquatic organisms at low water con-
centrations. Ability to concentrate high levels from water and reproductive
impairment in certain species are suggestive of potential environmental damage.
While the fate of these compounds remains obscure, their presence in water affects
successful growth and reproduction essential for maintenance of animal populations.
A freshwater criterion of 3 ug/1 is recommended even though some reproductive im-
pairment was seen In daphnids, since all other species tested were so much more
resistant. Until additional effect data become available this criterion should be
a goal for marine waters.
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REFERENCES CITED
Mayer, Jr., F.L., e_t al_., 1971. Phthalate esters: an environmental contaminant.
Presented at Midwest Regional A.C.S. meeting, Oct. 28-29, 1971,
St. Louis.
Mayer, Jr., F.L., et a\_., 1972. Phthalate esters as environmental contaminants.
Nature 238: 411.
Mayer, Jr., F.L. and H.O. Sanders, 1973. Toxicology of phthalate acid esters in
aquatic organisms. Environ. Health Perspec. 3:153.
Morris, R.J., 1970. Phthalate acid in the deep sea jellyfish, Atoll a.
Nature 227: 1264.
Sanders, H.O., et al_., 1973. Toxicity, residue dynamics and reproductive
effects of phthalate esters in aquatic invertebrates. Environ. Res. £: 84.
Stalling, D.C., 1972. Analysis of organochlorine residues in fish: Current
research at the Fish-Pesticide Research Laboratory. Pesticide Chemistry.
Vol. 4, Methods in Residue Analysis, A.S. Tahori, ed. (Gordon & Breach
Pub!., N.Y.), pp. 413.
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POLYCHLORINaTED BIPHENYIS
CRri'EKlON;
.001 ug/1 for freshwater and marine aquatic life
and for consumers thereof.
Every reasonable effort should be made to minimize human exposure.
INTRODUCTION;
Polychlorinatad biohenyls (PC3's) ara a class of
compounds produced by the chlorination of binhanyls and are
registered in the United States under the trade name,
Aroclor (R' . Tiie degree of chlorination determines their
chemical properties, and generally their composition cam be
identified oy the numerical nomenclature, e.g., \roclor
1242, \roclor 1254, etc. The first two digits represent the
molecular type and the last two digits the average
percentage oy weight of chlorine (NTIS, 1372).
Identification of PCB's in the presence of
organochlorine pesticides such as DDT and DDE has been
difficult in the past because of their similar
chromatographic characteristics (Risebrough, et aU, 1968).
-------
In PCB analysis today, the interference of organochloririe
hydrocarbons is overcome by sequential column chromatoqraphy
on Florisil and silica gel (Armour and Burke, 1970; FDA,
1971), Gas-liquid chromatography with highly sensitive and
selective detectors has been employed successfully in the
detection of PCB's at low levels (Nefceker and Puglisi,
1974).
PCB compounds are slightly soluble in water (25-200 ug/1
at 25°C), soluble in lipids, oils, organic solvents, and
resistant to both heat and biological degradation (NTIS,
1972; Nisbet, et al., 1972). Typically, the specific
gravity, boiling point, and melting point of PCB's increase
with their chlorine content. PCB's are relatively
non-flammable, have useful heat exchange and dielectric
properties, and now are used principally in the electrical
industry in capacitors and transformers.
RATIONALE;
The acute and chronic effects of PCB*s have been
determined on a number of aquatic organisms.
-------
Ninety-six-hour LC50 values for newly hatched fathead
minnows, Pimephales promelas, were 15 uq/1 for Aroclcr 12^2
and 7.7 uq/1 for Aroclor 1254. In 60-day continuous flow
bioassays 50 percent of the fathead minnows were killed in
8.8 uq/1 Aroclor 1242 and in 4,6 ug/1 Aroclor 1254 [Nebeker,
et aJL., 1974). Nine-month continuous flow bioassay tests
were conducted in the same series of experiments reported by
Nebeker, et al. (1974). The spawning of fathead minnows was
significantly affected at concentrations of 1,8 uq/1 Aroclor
1254; concentrations of Aroclor 1242 above 10 uq/1 were
lethal to newly hatched fry. Defoe, et al. (In Press)
conducted similar flow-through acute and chronic studies
with fathead minnows using Aroclor 1248 and 1260. The
calculated 30-day TL50 for newly hatched fathead minnows was
4.7 ug/1 for Aroclor 1248 and 3,3 ug/1 for Aroclor 1260.
Fathead minnows were able to reproduce successfully at PCB
concentrations which were acutely lethal to the larvae.
Stalling and Mayer (1972) determined 96-hour LC50 values
ranging from 1,170 to 50,000 ug/1 for cutthroat trout, Salmo
clarjci, using Aroclors 1221-1268. With rainbow trout, Salmo
qairdneri, and Aroclors 1242-1260 the acute toxicity was
-------
greater than 1500 mg/1. Fifteen—day intermittent-flow
bioassays carried out with bluegills, lepomis macrochirus,
and Arcclors 1242, 1248 and 1254 resulted in LC50 values of
54, 76 and 204 ug/1, respectively.
Johnson (1973), Mayer (1975) and Veith (1975) conducted
bioassays which showed that the toxicity of Aroclor 1016,
introduced recently to replace PCB's of the Aroclor 1200
series in many applications, was similar to that of Aroclor
1242.
Nebeker and Puglisi (1974) conducted bioassays with
paghnia magna exposed to concentrations of Aroclors
1221-1268. In continuous flow tests Aroclor 1254 was the
most toxic with a 3-week LC50 of 1.3 ug/1; 100 percent
mortality occurred at 3.5 ug/1 and 100 percent reproductive
impairment occurred at 3.8 ug/1. Stalling and Mayer (1972)
and Mayer et a_l. (1975) conducted flow-through and static
fcioassay tests on freshwater invertebrates which likewise
indicated that these organisms are generally more
susceptible to acute toxic effects of PCB's than fish.
-------
Studies of the Milwaukee River (Wisconsin) revealed PCS
concentrations in ambient water of 2.0 to 2.8 ug/1 and
residues in fish as high as 405 ug/g (Veith and Lee, 1971).
Open water Lake Michigan PCB concentrations have been
reported to be less than 0.01 ug/1; mean residues in coho
salmon, Oncorhynclrus kisutch, were about 15 ug/g (Veith,
1973) . Veith (1973) found that goldfish, Carassius auratus^,
in the lower Milwaukee River accumulated Aroclor 1248
approximately 0.7 x ID5 to 2 x 10 times depending upon the
ambient water concentration. From both laboratory and river
system studies, many aquatic organisms apoear to
bioaccumulate PCB mixtures containing 3, 4, 5, and 6
chlorine atoms per molecule approximately 3 x 10 to 2 x 10
times the concentration in the water. Tissue residues in
fathead minnows, Pimephales promelas, ranged from 0.7 ug/g
of Aroclor 1248 in control fish to 1036 ug/g of Aroclor 1254
in fish neld for 8 months in water containing 4.0 ug/1 of
Aroclor 1254 (Nebeker, et a.1., 1974). The latter case
indicates a bioaccumulation factor of 2.3 x 10 , which is
essentially independent of the PCS concentration in the
water.
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6
Bluegill sunfish, Lepomis macrochirus, exposed to
Aroclors 1248 and 1254 exhibited a bioaccumulation factor of
7.1 x 104 (Stalling and Mayer, 1972). The bioaccumulation
factor for gizzard shad, Dorosoma cepedianum, in the Saginaw
River (Michigan) varied between 0.6 x 105 and 1.5 x 10 for
Aroclor 1254 (Michigan v«/ater Resource Commission, 1973) .
On the basis of the FDA action level of 5 ug/g in fish
tissue, and both laboratory and field derived
bioaccumulation levels for fathead minnows, Pimephales
promelas, and goldfish, Carassius auratus, in the range of
3 5
0.7 x 10 to 2.3 x 10 , an ambient water concentration of no
more than 0.022 ug/1 would be permissible. However, lake
trout from the Great Lakes larger than.12 inches in length
generally contain more than 5 ug/g of PCB's and chub from
the Great Lakes generally contain PC3's in amounts
approacning or slightly exceeding 5 ug/g. Since monitoring
data on the Great Lakes' waters consistently indicate
concentrations equal to or less than .01 ug/1, a criterion
of less than 0.01 ug/1 for all fishes appears necessary.
-------
A residue level of 2 ug/g in fish consumed
-------
8
pathways and effects of PCB's in the estuarine and marine
environments. The sediment reservoir of Aroclor 1254 is
thought to be a continuing source of PCB's to aquatic biota.
The initial survey of Escambia Bay biota revealed fish,
shrimp, and crabs with levels as high as 12 ug/g. Higher
levels of PCB's were detected in higher trophic levels than
shrimp, which could implicate a chain transfer from sediment
to large animals (Duke, 1974).
From the Escambia Bay data, which include flow-through
bioassays with residue analyses where possible, the
following conclusions were reached: (1) all of the Aroclcrs
tested are acutely toxic to certain estuarine organisms; (2)
bioassays lasting longer than 96 hours demonstrated that
Aroclor 1251 is toxic to commercial shrimp at less than 1
ug/1; (3) fish, particularly sheepshead minnows, Cyprinodon
variegatus, are extremely sensitive to Aroclor 1254 with 0.1
ug/1 lethal to fry; and, (4) acute toxicity of Aroclor 1016
to estuarine organisms is similar to that of other Aroclors
but it appears less toxic to marine fish in long-term
exposures than does Aroclor 1254 (Duke, 1974; Schimmel, et
al., 1974).
371
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Oysters, Crassgstrea virginlca, were sensitive-to-
Aroclor 1260 with growth diminished by 44 percent in
concentrations of 10 uq/1 and by 52 percent in 100 tig/1.
Approximately 10 percent of the pink shrimp, Penaeus
duorarum, died in 100 uq/1, but no apparent effects on
pinfish, Lagodon rhornboides, were noted at that
concentration. Aroclor 1254 had no apparent effect on
juvenile pinfish at 100 uq/1 in 48-hour flow-through tests,
but killed 100 percent of the pink shrimp. At. 100 ug/1 of
Aroclor 1254 for 96* hours, shell growth of oysters was
^inhibited and was decreased 41 percent at levels of 10 ug/1.
The toxicity of Aroclor 1248 and 1242 to shrimp arid pinfish
was similar to that of Aroclor 1254. Aroclor 1242 was toxic
to oysters at 100 ug/1. Killifish, Fundulus heteroclitus,
exposed to 25 ug/1 of Aroclor 1221 suffered 85 percent
mortality. in 96-hour bioassays, Aroclor 1016 was toxic to
an estimated 50 percent of the oysters, crassostrea
virqinica; brown shrimp, Penaeus aztecus; and grass shrimp,
Palecmonetes pugig, at 10 ug/1; it was lethally toxic to 18
percent of the pinfish at 100 ug/1 (Duke, 1974).
372
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10
Younq oysters, CrassQStrea virqinica, exposed to Aroclor
1254 in flowing sea water for 24 weeks experienced reduction
in qrowt.h rates at 4.0 uq/1, but apparently were not
affected by 1.0 ug/1. Oysters accumulated as much as
100,000 times the test-water concentration of 1.0 uq/1.
General tissue alterations in the vesicular connective
tissue around the .diverticula of the hepatoparicreas were
noted in the oysters exposed to 5,0 uq/1. No significant
mortality was observed in oysters exposed continuously to
0,01 uq/1 of Aroclor 1254 for 56 weeks (Duke, 1974).
Blue crabs, Call in ect.es sapidus, apparently were not
affected by 20 days' exposure to 5.0 uq/1 of Aroclor 1254.
Pink shrimp exposed under similar conditions experienced a
72 percent mortality. In subsequent flow-through bioassays,
51 percent of the -juvenile shrimp were killed by Aroclor
1254 in 15 days and 50 percent of the adult shrimp were
killed at 3.0 uq/1 in 15 days. From pathological
examinations of the exposed pink shrimp, it appears that
Aroclor 1254 facilitates or enhances the susceptibility to
latent viral infections. Aroclor 1254 was lethal to qrass
shrimp, Paleomonetes gugio, at 4.0 ug/1 in 16 days, to
373
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11
amphipods at 10 ug/1 in 30 days, and to juvenile spot,
Leioatontua xanthurus, at 5.0 ug/1 after 20 to 45 days.
Shespshead minnows, Cyprinodon variegatus, were tha most
sensitive estuarine organisms to Aroclor 1254 with 0.3 ug/1
lethal to the fry within 2 weeks. Aroclor 1016, in two
different 42-day flow-through bioassays, caused significant
mortalities of pinfish at 32 ug/1 and 21 ug/1. Pathological
examination of those exposed to 32 ug/1 revealed several
liver and pancreatic alterations. Sheeoshead minnows in
28-day \roclor 1016 flow-through bioassays were not affected
by concentrations of 10 ug/1 or less, but died at 32 and 100
ug/1 (Duke, 1974) . The bioaccuraulation factors determined
by the flow-through bioassays are:
Aroclor
1254
1254
1254
1254
Accumulation Factors
(as a multiplier of test
Organism
Oyster (Crassostrea
virginica) "~
3lue crab
(Callinectes s_a_pidu3)
Grass shrimp
(Paleomonetes pugicO
Spot (Leiostomus
Time
30 days
20 days
7 days
14-28 days
water concentrations)
1.01 x 105
4 x 103
3.2 x 103 to
37 x 1Q3
11 x 103
xanthurus)
-------
12
1254 Pinfis-i 35 days 21.8 x 103
(Lagodon rhomboides)
1016 Pinfish 42 days 11 x 103 to 24 x 103
(Lagodon rhomboides)
1316 Sheepshead minnow 2.5 x 103 to 8.1 x 103
(Cyprinodon variegatus)
3asad upon an accumulation factor of 100,000 in the
oyster, it is necessary to limit the marine water
concentration of PC3's to a maximum of 0.01 ug/1 to protect
the human consumer. However, data on the toxicity of
Aroclor 1254 to sheepshead minnow fry mentioned earlier
(Schimmel, et al., 1974), which indicate lethality at 0.1
ug/1, justify lowering the latter concentration by a factor
of 0.01 to obtain a marina criterion of 0.001 ug/1. This
level is furtner supported by the evidence cited earlier
suggesting that a food tissue level of 0.5 ug/g, or 0.1
times the FDA level for human consumption, is necessary to
protect carnivorous mammals.
Evidence is accumulating that PCB's do not contribute to
shell tninning of bird eggs (N^S, 1974). Dietary PCB's
produced no shell thinning in eggs of mallard ducks (Heath,
at al., 1972). PCB's may increase susceptibility to
-------
13
infectious agents such as viral diseases (Friend and
Trainer, 1970), and increase the activity of liver enzymes
that degrade steroids, including sex hormones (Risebrough,
et al., 1968; Street, et al., 1968). Laboratory studies
have indicated that PCB's with their derivatives or
metabolites, cause embryonic death of birds (Voss and
Koeman, 1970). Feeding PCB's to White Leghorn pullets at a
level of 20 ppm caused a significant decrease in
hatchability of the eggs and viability of the surviving
progeny (Lillie, et al., 197U; Lillie, et al., 1975); in
many cases, the cause of embryo mortality was attributed to
gross abnormalities which ranged from edema to malformed
s
appendages (Cecil, et al., 197U).
Exposure to PCB's is known to cause skin lesions
(Schwartz and Peck, 19 U3) and to increase liver enzyme
activity that may have a secondary effect on reproductive
processes (Risebrough, 1969; Street, et al., 1968;
Wasserman, et al., 1970). It is not clear whether the
effects are due to the PCB's or their contaminants, the
chlorinated ditenzofurans, which are highly toxic (Bauer, et
al.f 1961; Schultz, 1968; Varrett, 1970). While chlorinated
-------
14
dibenzofurans are a by-product of PCB production, it is not
known whether they are also produced by the degradation of
PCB's (NAS, 1974).
377
-------
15
REFERENCES CITED:
Armour, J.A. and J.A. Burke, 1970. Method for separating
polychlorinated biphenyls from DDT and its analogs.
Assoc. Official Anal. Chem,, 53:761.
Bauer, H., et al., 1961. Occupational poisonings in the
production of chlorophenol compounds. Arch.
Gewerbepath, 18:583.
Cecil, H.C., et al., 1974. Embryotoxic and teratogenic
effects in unhatched fertile eggs from hens fed
polychlorinated biphenyls (PCB's). Bull. Envir. Contam,
and Tox. , 11:489,
DeFoe, D.L., eit al_.» 1976. Effects
cf Aroclor 1248 and 1260 on the fathead minnow,. J.
Fish. Res. Board Canada (In Press).
Duke, T.W., 1974. Testimony in the matter of proposed toxic
pollutant effluent standards for Aldrin-Dieldrin et al.
FWPCA (307) Docket #1.
-------
16
Food and Drug Administation, 1971. U.S. Dept. of H.E.W.
Pesticide Analytical Manual, Volumes I and II.
Friend, H. and D.O. Trainer, 1970. Polychlorinated biphenyl
interaction with duck hepatitis virus. Science,
170:1314.
Hays, H. and R.W. Risebrough, 1972. Pollutant
concentrations in abnormal spring tern from Long Island
Sound. AUK, 89:19.
Heath, R.G., et al., 1972. Effects of polychlorinated
biphenyls on birds. Proceedings IV International
Ornithological Congress (In Press).
Johnson, W., 1973. Unpublished report on toxicity of
Aroclor 1016. Fish-Pesticide Laboratory, USDI, Fish and
Wildlife service, Columbia, MO 65201.
Lillie, R.J., et al., 1974. Differences in response of
caged white Leghorn layers to various polychlorinated
biphenyls (PCB's) in the diet. Poultry Science, 53:726.
379
-------
17
Lillie, R.J., et £1., 1975. Toxicity of certain
polychlorinated and polybrominated biphenyls on
reproductive efficiency of caged chickens. Poultry
Science, 54:1559.
Mayer, F.L., fii s!'• 1975. Residue
dynamics and biological effects of PCB's in aquatic
organisms. Arch. Env. Cont. and Toxic. (In Press).
Mayer, F.L. 1975. Toxicity of Aroclor 1016 to freshwater
fishes. Unpublished data. Fish-Pesticide Laboratory,
USDI, Fish and Wildlife Service, Columbia, MO 65201.
Michigan Water Resource Commission, 1973. Monitoring for
polychlorinated biphenyls in the aquatic environment.
National Academy of Sciences, National Academy of
Engineering, 1974. Water quality criteria, 1972. U.S.
Government Printing Office, Washington, D.C.
380
-------
18
National Technical Information Service, 1972.
Polychlorinated fciphenyls and the environment. NTIS
Interdepartmental Task Force, Washington, D.C.
Nebeker, A.V., and F.A. Pugl^si, 197U. Effect of
polychlorinated biphenyls (PCB«s) on survival and
reproduction of Daphnia, Gammarus, and Tanytarsus.
Trans. Amer. Fish. Soc. 103 (U): 722.
Nebeker, A.V., et al_., 197U. Effect
of polychlorinated biphenyl compounds on survival and
reproduction of the fathead minnow and flagfish. Trans.
Amer. Fish. Soc. 103(3): 562.
Nisbet et al., 1972. Rates and routes of transport of PCB's in the
environment. Environ. Health Perspectives 4:21.
Platonow, N.S. and L.H. Kalstad, 1973. Dietary effects of
polychlorinated biphenyls en mink. Can. Jour. Comp.
Med., 37:391.
Ringer, P.K., et aj.., 1972. Effect of dietary
polychlorinated biphenyls on growth and reproduction of
-------
19
mink. Amer. Chem. Soc,, National Meeting Reprints of
Papers, 12:1U9,
Risebrough, R.W., et a1., 1968, Polychlorinated biphonyls
in the qlobal ecosystem. Nature, 220:1098.
Risebrough, R.W., 1969. Chlorinated hydrocarbons in marine
ecosystems. Iff: Chemical Fallout, M.fo. Miller and G.G,
Berq, eds., C.C. Thomas, Springfield,
Risehrough, R.W., 1970. More letters in the world.
Environ., 12:16.
Schimmc-l, s.c., et al. , 197U. Effects of Aroclor 125U on
laboratory-reared embryos and fry of sheepshead minnows
(Cyprinodon variegatus). Trans. Amer. Fish. Soc,,
103:582.
Schultz, K.H,, 1968. Clinical picture and etiology of
chloracne. Arbeitsmed. Sozialtned. Arbeitshgy, 3:25.
Reprinted in translation in U.S., Congress. Senate,
Committee on Commerce, Effects of 2, U, 5-T on man and
-------
20
the environment: Hearings, 91st Cong., 2 sess., pp,
336-341.
Schwartz, L. and s.M. Peck, 1913. occupational acne. New
York State Med., 13:1711.
Stalling, D.L. and F,L. Mayer, Jr., 1972. Toxicities of
PCfi's to fish and environmental residues. Env. Health
Perspectives, 1:51.
Street, j.c., et al., 1968. Comparative effects of
polychlorinated biphenyls and organochlorine pesticides
in induction of hepatic microsomal enzymes. Aroer, Chew.
Soc., 15flth National Meeting. Sept. 8-12, 1968.
Varrett, J,, 1970. Statement in n.s. Congress, Senate,
Committee on Commerce, Effects of 2, 1, 5-T on wan and
the environment: Hearings, 91st Cong,, 2nd sess., pp,
190-360.
Veith, G.D., 1973. Chlorinated hydrocarbons in fish and
lake fish. Final report, EPA Grant 1602 PBE.
383
-------
21
TQ71 chlorobiphenyls (PCE's) in
Veith, G.D. and G.F. Lee, 1971. cniorc. F
the Milwaukee Piver, Water Pes. , 5:1107,
V«ith, G.D. 1975. Personal communication. PCB test data,
National Water Quality Laboratory. Duluth, MN, ^580^.
Voss, J.G. and J.H. Koeman, 1970, Comparative t.oxicologic
stu^y with polychlorinat^ biphenyls in chickens with
special refine* to pbrphyria, edema formation, liver
nephrosis, and tissue residues, Toxicol. Apl.
Pharmacoi. , 17:656.
i la-rft TV.O pfffr-t of orqanochlorine
Wasserman. M., et aj.. , 1970. The ettecr H
insecticides on serum cholesterol level in people
occupational^ exposed. Bull. Fnviron. Contam.
Toxicol., 5:368.
3SH
-------
SELENIUM
CRITERIA:
10 ug/1 for domestic water supply (health);
For marine and freshwater aquatic life; 0.01 of
the 96-hour LC§0 as determined through bioassay
using a sensitive resident species.
INTRODUCTION:
Biologically, selenium is an essential, beneficial element recognized
as a metabolic requirement in trace amounts for animals but toxic to them
when ingested in amounts ranging from about 0.1 to 10 mg/kg of food. The
national levels of selenium in water are proportional to the selenium in
the soil. In low selenium areas, the content of water may be well below
1 ug/1 (Lindberg, 1968). In water from seleniferous areas, levels of
selenium of 50 to 300 ug/1 have been reported (WHO, 1972).. Selenium
appears in the soil as basic ferric selenite, calcium selenate, and as
elemental selenium. Elemental selenium must be oxidized to selenite or
selenate before it has appreciable solubility in water.
RATIONALE;
Selenium is considered toxic to man. Symptoms appear similar to
those of arsenic poisoning (Keboe, et al_., 1944; Fairhill, 1941). Any
consideration of the toxicity of selenium to man must take into con-
sideration the dietary requirement for the element in amounts estimated
to be 0.04 to .10 mg/kg of food. Considering this requirement in
conjunction with evidence that ingestion of selenium in amounts as low
385
-------
as .07 mg per .day has been shown to give rise to signs of selenium
toxicity, selenium concentrations above 10 ug/1 should not be permitted
in drinking water (Smith, et^ al_., 1936; Smith and West, 1937). The
USPHS drinking water standards recommend that drinking water supplies
contain no more than 0.01 mg/1 of selenium (USPHS, 1962).
As sodium selenite, 2.0 mg/1 of selenium has been demonstrated to
be lethal to goldfish, Carassius auratus, in 18 to 46 days (Ohio River
Valley Water Sanitation Commission, 1950). Bringmann and Kuhn (1959)
demonstrated the threshold effect of selenium as sodium selenite on
a freshwater crustacean, an alga and a bacterium. In two days the
median threshold effect occurred at 2.5 mg/1 with Daphnia; in 4 days the
median threshold effect was 2.5 mg/1 with Scenedesmus, at 90 mg/1 with
Escherichia coli:, and 183 mg/1 for the protozoan, Microregma..
Selenium in water apparently is toxic at concentrations of 2.5 mg/1
or less to those few species tested. Animals can beneficially metabolize
ingested selenium in amounts of 0.01 to 0.10 mgAg of food.
Based on the data available, freshwater fish should not he exposed
to water containing more than 0.01 of the 96-hour I£50 as determined
through bioassay using a sensitive resident species.
-------
REFERENCES CITED:
Bringmann, G. and R. Kuhn, 1959. The toxic effect of wastewater on aquatic
bacteria, algae and small crustaceans. Gesundeits-Ing., 80:115.
Fairhill, L.T., 1941. Toxic contaminants of drinking water. Jour. North
East Water Works Assn., 55:400.
Keboe, R.H., ejt al_., 1944. The hygienic significance of the contamination
of water with certain mineral constituents. Jour. Amer. Water Works
Assn., 36:645.
Lindberg, P., 1968. Selenium determination in plant and animal material
in water. Acta Vet Scand., Suppl. 32, Stockholm.
Ohio River Valley Water Sanitation Commission, 1950. Subcommittee on
toxicities-metal finishing. Industries Action Committee, Report No. 3.
Smith, M.I., e_t al_., 1936. The selenium problem in relation to public
health. Public Health Report 51, (U.S.) 51:1496.
Smith, M.I. and B.B. West, 1937. Further field studies on the selenium
problem in relation to public health. Public Health Report, 52:1375.
U.S. Public Health Service, 1962. Drinking water standards, 27 F.R. 2152.
World Health Organization, 1972. Health hazards of the human environment.
Geneva.
-------
SILVER
CRITERIA;
50 ug/1 for domestic water supply (health).
For marine and fresh water aquatic life 0. 01
of the 96-hour LC CJQ as determined through bio-
assay using a sensitive resident species.
INTRODUCTION;
Biologically, silver is a non-essential, non-beneficial element
recognized as causing localized skin discoloration in humans, and as
being systemically toxic to aquatic life. Digestion of silver or silver
salts by humans results in deposition of silver in skin, eyes and mucous
membranes that causes a blue-gray discoloration without apparent
systemic reaction (Hill, 1957). Because of its strong bactericidal action,
silver has been considered for use as a water disinfectant. Dosages of
0.001 to 500 ug/1 of silver have been reported sufficient to sterilize
water (McKee and Wolf, 1963). At these concentrations, the ingestion of
silver has no obvious detrimental effect on humans.
RATIONALE;
The 1962 USPHS Drinking Water Standards contained a limit for silver
of 0.05 mg/1. This limit was established because of the evidence that
silver, once absorbed, is held indefinitely in tissues, particularly the
skin, without evident loss through usual channels of elimination or re-
duction by transmigration to other body sites, and because of the probable
high absorbability of silver bound to sulfur components of food cooked in
silver-containing water (Aub and Fairhall, 1942). A study of the toxic
-------
effects of silver added to drinking water of rats showed pathologic
changes in kidneys, liver and spleen at concentrations of 400, 700, and
1, 000 ug/1 (Just and Szniolis, 1936).
Using silver nitrate, Coleman and Clearly (1974) demonstrated that
juvenile largemouth bass, Micropterus salmoides, and bluegill,
Lepomis macrochirus exposed to silver in concentrations of 0. 3 to
70 ug/1 accumulated the metal, especially in the internal organs and
gills. The quantity of accumulated silver increased as exposure
concentrations increased with a subsequent equilibrium developing
between the water and tissue concentrations. After two months'
exposure to 7 ug/1 silver, the concentration of silver in gills
exceeded that in the gills of the control fish by 200-fold. The 70 ug/1
concentration of silver was 1ethal1y toxic to bass.
Data compiled by Doudoroff and Katz (1953) show that sticklebacks
were killed by a 20 ug/1 concentration of silver nitrate in two days.
Anderson (1948) reported that the toxic threshold of silver nitrate
for sticklebacks, Gasterosteus aculeatus, was 3.0 ug/1 as silver.
He also found that Daphnia magna were immobilized by 3.2 ug/1 as
silver. Jones (1939) found the lethal concentration of silver nitrate
for sticklebacks was 3.0 ug/1 as silver. In differing concentrations
the average survival times of the fish were; one week at 4. 0 ug/1,
four days at 10. 0 ug/1, and only one day at 100. 0 ug/1.
-------
In a 10-month bioassay on rainbow trout, Salmo gairdneri.
exposed to silver nitrate, Geottl, et^al. (1974) determined that no
significant number of test fish died when exposed to silver
concentrations of . 09 and 0.17 ug/1 as silver. The results did not
reflect possible effects of silver on spawning behavior or reproduction.
Using silver thiosulfate, silver nitrate, silver carbonate and
silver chloride, Terhaar, et^al. (1972) reported that all of a test
group of 20 fathead minnows, Pimephales promelas. survived
exposure for 96 hours to 5,000 ug/1 silver as silver thiosulfate;
at 250,000 ug/1 15 of 20 fish died. Silver nitrate 1n
concentrations of 1,000 and 100.0 ug/1 3S silver killed 16 out of
20 and 12 out of 20 fish, respectively. Silver carbonate killed all
20 test fish at concentrations of 1,000 ug/1.
In marine waters a concentration of 400 ug/1 as silver killed 90
percent of test barnacles, Balanus balanoides (Clarke, 1947).
Silver nitrate effects on the development of sea urchin, Arbacia,
have been reported at approximately 0. 5 ug/1 (Wilber, 1969).
Calefcrese, et al. (1973) reported an LC (48-hour) of 5. 8 ug/1 as silver
50
for oyster larvae, Crassostrea virginica, and an LC (48-hour) of
«2 5Q
21.0ug/J as silver for larvae of the hard-shell clam, Mercenaria
mercenaria. Jackim, et al. (1970) reported a sublethal enz3rme
effect at a concentration of 20.0UQ/1 as silver for Fundulus heteroclltus.
390
-------
It is apparent that there is a wide variation in the toxicity of silver
compounds to aquatic life and that the degree of -dissociation characteristic
of these compounds affects toxicity. Since little information is available
on the movement and chemical stability of these compounds in the
aquatic environment, a silver criterion must be based on the total
silver concentration.
The silver criterion should be established at 0.01 of the 96-hour
LC50 aS determined through bioassay using a sensitive resident species.
This application factor has been recommended (NAS , 19?A for
persistent or cumulative toxicants.
391
-------
REFERENCES CITED;
Anderson, B. G., 1948. The apparent thresholds of toxicity of
Daphnia magna for chlorides of various metals when added to
Lake Erie water. Trans. Amer. Fish. Soc., 78-96.
Aub, J. C. and L. T. Fairhall, 1942. Excretion of silver in urine.
Jour. Amer. Medical Assn., 118-316.
Clarke, G. L., 1947. Poisoning and recovery in barnacles and
mussels. Biol. Bull., 92:73.
Calabrese> A.., et^al., 1973. The toxicity of heavy metals to
embryos of the American oyster Crassostrea virginica.
Marine Biology, 18:162.
Coleman, R. L. and J. E. Clearly, 1974. Silver toxicity and
accumulation in largemouth bass and bluegill. Bull, of
Environmental Contamination and Toxicology, Vol. 12 No. 1.
Doudoroff, P. and M. Katz, 1953. Critical review of literature of
toxicity of industrial wastes and their components to fish.
Sew. and Ind. Wastes, 25:802.
Geottl, J. P., et^aL. 1974. Water Pollution studies. Job Progress
Report, Colorado Department of Natural Resources, Federal Aid
Project F-33-R-9 41-44.
Hill, W. B., e^aU , 1957. Argyria investigation- -toxicological
properties of silver. Amer. Silver Producers. Res. Proj. Rept.,
Appendix II.
-------
Jackim, E., et al., 1970. Effects of metal poisoning of five liver
enzymes in killifish (Fundulus heteroclitus). Jour. Fish. Res.
Bd. Can., 27:388.
Jones, J. R. E., 1939. The relation between the electrolytic solution
pressures of the metals and their toxicity to the stickleback. Jour.
Exp. Biol., 16:425.
Just, J. and A. Szniolis, 1936. Germicidal properties of silver in
water. Jour. Amer. Water Works Assn., 28:492.
McKee, J.E. andH.W. Wolf, 1963. Water quality criteria. California
State Water Resources Control Board, Publ. No. 3-A.
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972. U. S. Government Printing Office,
Washington, D. C.
Terhaar, C.J., et^ah , 1972. Toxicity of photographic processing
chemicals to fish. Photo. Sci. andEng., 16:370.
Wilber, C. G., 1969. The biological aspects of water pollution. Charles
Thomas, Springfield, Illinois, p. 296.
393
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SOLIDS (DISSOLVED) AND SALINITY
CRITERION:
250 mg/1 for chlorides and sulfates
in domestic water supplies ^welfare).
INTRODUCTION;
Dissolved solids and total dissolved solids are terms generally associated
with freshwater systems and consist of inorganic salts, small amounts of organic
matter and dissolved materials (Sawyer, 1960). The equivalent terminology -in
Standard Methods is filtrable residue (Standard Methods, 1971). Salinity is
an oceanographic term, and although not precisely equivalent to the total
dissolved salt content it is related to it (Capurro, 1970). For most purposes,
the terms total dissolved salt content and salinity are equivalent. The principal
Inorganic anions dissolved in water include the carbonates, chlorides, sulfates
and nitrates (principally in ground waters); the principal cations are sodium,
potassium, calcium, and magnesium.
RATIONALE:
Excess dissolved solids are objectionable in drinking water because of
possible physiological effects, unpalatable mineral tastes and higher costs
because of corrosion or the necessity for additional treatment.
The physiological effects directly related to dissolved solids include
«
laxative effects principally from sodium sulfate and magnesium sulfate and the
adverse effect of sodium on certain patients afflicted with cardiac
disease and women with toxemia associated with pregnancy. One study was made
394
-------
using data 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/1, 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/1 (Moore, 1952). To
protect transients to an area, a sulfate level of 250 mg/1 should afford
reasonable protection from laxative effects.
As indicated, sodium frequently is the principal component of dissolved solids.
Persons on restricted sodium diets may have an intake restricted from 500 to
1,000 mg/day (Nat. Res. Coun., 1954). That portion ingested in water must be
compensated by reduced levels in food ingested so that the total does not
exceed the allowable intake. Using certain assumptions of water intake
(e.g., 2 liters of water consumed per day) and sodium content of food, it has
been calculated that for very restricted sodium diets, 20 mg/1 in water would be
the maximum, while for moderately restricted diets, 270 mg/1 would be maximum.
Specific sodium levels for entire water supplies have not been recommended but
various restricted sodium intakes are recommended because: (1) the general
population is not adversely affected by sodium, but various restricted sodium
intakes are recommended by physicians for a significant portion of the population,
and; (2) 270 mg/1 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 water supplies is costly (NAS, 1974).
A study based on consumer surveys in 29 California water systems was made to
measure the taste threshold of dissolved salts in water (Bruvold eit aj_., 1969)
t
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/1 dissolved solids as "excellent" while those with 1283 to 395"
-------
1333 mg/1 dissolved solids were "unacceptable" depending on the rating system
used. A "good" rating was registered for dissolved solids less than 658 to 755 mg/1.
These results should be interpreted with consideration of consumer acclimation to
such waters. The PHS Drinking Water Standards (1962) recommended a maximum dissolved
solids concentration of 500 mg/1 unless more suitable supplies were unavailable.
Specific constituents included in the dissolved solids in water may cause
mineral taste at lower concentrations than other constituents. Chloride Ions
have frequently been cited as having a low taste threshold in water. Data
from Rlcter and MacLean (1939) on a taste panel of 53 adults indicated that.
61 mg/1 NaCl was the median level for detecting a difference from distilled
water. At a median concentration of 395 mg/1 chloride a salty taste was distinguish-
able, although the range was from 120 to 1215 mg/1. Lockhart, et al_. (1955)
evaluated the effect of chlorides on water used for brewing coffee. Threshold
concentrations for chloride ranged from 210 mg/1 to 310 mg/1 depending on the
associated cation. These data indicate that a level of 250 mg/1 chlorides is a
reasonable maximum level to protect consumers of drinking water.
The causation of corrosion and encrustation of metallic surfaces by water
containing dissolved solids is well known. In water distribution systems
corrosion is controlled by insulating dissimilar metal connections by non-metallic
materials, use of pH control and corrosion inhibitors, or some form of galvanic
or impressed electrical current systems (Lehmann, 1964). Damage in household
systems occurs to water piping, wastewater piping, water heaters, faucets, toilet
T
flushing mechanisms, garbage grinders and both clothes and dishwashing machines.
-------
By use of water with 1,750 mg/1 dissolved solids as compared with 250 mg/1,
service life reductions ranged from 70 percent for toilet flushing mechanisms
to 30 percent for washing equipment. Such increased corrosion was calculated
in 1968 to cost the consumer an additional $0.50 per 1,000 gallons used
(Patterson and Banker, 1968).
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/1 dissolved solids, that whitefish and
pike-perch survived 15,000 mg/1, but only the stickleback survived 20,000 mg/1
dissolved solids. It was concluded that lakes with dissolved solids in excess
of 15,000 mg/1 were unsuitable for most freshwater fishes (Rawson and Moore,
1944). The NTAC Report (1968) also recommended maintaining osmotic pressure
levels of less than that caused by a 15,000 mg/1.solution of sodium chloride.
Marine fishes also exhibit variance in ability to tolerate salinity
changes. However, fishkills in Laguna Madre off the Texas coast have occurred
with salinities 1n the range of 75 to 100 o/co. Such
concentrated sea water 1s caused by evaporation and lack of exchange with the
Gulf .of Mexico (Rounsefell and Everhart, 1953).
Estuarine species of fish are tolerant of salinity changes ranging
from fresh to brackish to sea water. 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
397
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certain minimal salinities for successful hatching of eggs. The oyster drill
cannot tolerate salinities less than 12.5 o/oo. Therefore, estuarine segments con-
taining salinities below about 12.5 o/oo produce most of the seed oysters for planting
(Rounsefell and Everhart, 1953). Based on similar examples, the NTAC Report
(1968) 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 v/ater have been extrapolated from data collected on domestic species
such as chickens. However, young ducklings were reported poisoned in Suisan
Marsh by salt when maximum summer salinities varied from 0.55 to 1.74 o/oo with means
as high as 1.26 o/oo (Griffith, 1963).
Indirect effects of excess dissolved solids are primarily the elimination of
desirable food plants and other habitat forming plants. Rapid salinity changes
cause plasmolysis of tender leaves and stems because of changes in osmotic
pressure. The NTAC Report (1968) recommended the following limits in salinity
variation from natural to protect wildlife habitats:
Natural Salinity Variation Permitted
(o/oo) (o/op)
0 to 3.5 1
3.5 to 13.5 2
13.5 to 35 4
398
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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/1 of salts of sodium and calcium
combined with bicarbonates, chlorides and sulfates but only 10,000 mg/1 of
corresponding salts of potassium and magnesium. The approximate limit for highly
alkaline waters containing sodium and calcium carbonates is 5,000 mg/1 (NTAC, 1968),
Irrigation use of water is not only dependent 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).
Dissolved Solids Hazard for Irrigation Hater (mg/1)
Jfable 9 a ' * *' '
Water from which no detri-
mental effects will usually
be noticed 500
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,
399
-------
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 reaarding the dissolved solids content of raw
waters is quite variable. Table. 10indicates 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.
*ESP = 100[a + b(SAR)]
1 [a + b(SAR)l
where: a = intercept representing 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
SAR is expressed as milliequivalents
-------
Table 10
Total Dissolved Solids Concentrations of Surface
Waters that have been Used as Sources for
Industrial Water Supplies
Industry/Use Maximum Concentration
(mg/1)
Textile 150
Pulp and Paper 1,080
Chemical 2,500
Petroleum 3,500
Primary Metals 1,500
Copper Mining 2,100
Boiler Make-up 35,000
Source: NAS, 1974
-------
REFERENCES CITED:
Agriculture Handbook No. 60, 1954. Diagnosis and improvement of saline
and alkali soils. L. A. Richards, editor, U.S. Government Printing Office,
Washington, D. C.
Bruvold, W.H., ert al_.» 1969. Consumer assessment of mineral taste in domestic
water. Oour. Amer. Water Works Assn., 61;575.
Capurro, L.R.A., 1970. Oceanography for practicing engineers. Barnes
and Noble Inc., New York.
Griffith, E.W., 1963. Salt as a possible limiting factor to the Suisan
Marsh pheasant population. Annual Report Belta Fish and Wildlife
Protection Study, Cooperative Study of California
Lehman, J.A., 1964. Control of corrosion in water systems. Oour. Ameie
Water Works Assn., 56: 1009.
Lockhart, E.E., e_t al_., 1955. The effect of water impurities on the flavor
of brewed coffee. Food Research, 20:598.
Moore, E.W., 1952. Physiological effects of the consumption of saline drinking
water. National Res. Council, Div. of Medical Sciences, Bull. San. Engr.,
and Environment, Appendix E.
National Academy of Sciences, National Acadary of Engineering, 1974.
Water quality criteria, 1972. U.S. Government Printing Office, Washington, D.C.
National Research Council, 1954. Sodium restricted diets. Publication 325,
Food and Nutrition Board, Washington, D.C.
-------
National Technical Advisory Committee to the Secretary of thellftterior, 1968.
Water quality criteria. U.S. Government Printing Office, Washington, D.C.
Patterson, W.L. and R.F. Banker, 1968. Effects of highly mineralized water on
household plumbing and appliances. Jour. Amer. Water Works Assn., 60:1060.
Public Health Service, 1962. Drinking water standards, 1962. U.S. Govt.
Print. Office, Washington, D. C.
Rawson, D.S. and J,E. Moore, 1944. The saline lakes of Saskatchewan.
Canadian Jour, of Res., 22: 141.
Ricter, C.O., and A. Maclean, 1939. Salt taste threshold of humans.
Am. J. Physiol., 126:1.
Rounsefell, G.A. and W.H. Everhart, 1953. Fishery science, Its methods and
applications. John Wiley and Sons, Inc., New York.
Sawyer, C.N., 1960. Chemistry for sanitary engineers. McGraw-Hill Book
Co., Inc., New York.
Standard methods for the examination of water and wastewater, 13th edition,
1971. APHA, Amer. Water Works Assn., Water Poll. Control Fed.,
Washington, 0. C.
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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 photosythetic
activity by more then 10 percent from the seasonably
established norm 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 micro-
organisms do not come Into contact with the chlorine disinfectant (NAS, 1974).
The ability of common water treatment processes (i.e., coagulation, sedimentation,
filtration and chlorinatlon) to remove suspended matter to achieve acceptable
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final turbidities 1s a function of the composition of the material as well as
Its concentration. Because of the variability of such removal efficiency, 1t 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 of the possibility of unseen submerged hazards and the difficulty
1n locating swimmers 1n danger of drowning (NAS, 197*). 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.
F1sh. 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 identified four
effects on the fish and fish food populations, namely:
"(1) by acting directly on the fish swimming in water in which
sol Ids are suspended, and either killing them or reducing their growth
rate, resistance to disease, etc.;
(2) by preventing the successful development of fish eggs and larvae;
(>3) by modifying natural movements and migrations of fish;
(4) by reducing the abundance of food available to the fish; . . ."
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Settleable materials which blanket the bottom of water bodies damage the
Invertebrate populations, block gravel spawning beds, and 1f organic, remove
dissolved oxygen from overlying waters (EIFAC, 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/1, 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 concentrations
(Gammon, 1970). Similar effects have been reported downstream from an area
which was intensively logged. Major increases in stream suspended solids
(25 ppm turbidity upstream vs. 390 ppm downstream) caused smothering of bottom
Invertebrates reducing organism density to only 7.3 per square foot versus
25.5 per square foot upstream (Tebo, 1955).
When settleable sol Ids block gravel spawning beds which contain eggs,
high mortalities result although there is evidence that some species of salmonids
will not spawn in such areas (EIFAC, 1965).
It has been postulated that silt attached to the eggs prevents sufficient
exchange of oxygen and carbon dioxide between the egg and the overlying water.
The Important variables are particle size, stream velocity and degree of turbulence
(EIFAC, 1965).
Deposition of organic materials to the bottom sediments can cause imbalances
In stream biota by increasing bottom animal density, principally worm populations,
and diversity is reduced as pollution sensitive forms disappear (Mackenthun,
1973). Algae likewise fluorish in such nutrient rich areas although forms
may become less desirable (Tarzwell and Gaufin, 1953).
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Plankton and Inorganic suspended materials reduce light penetration Into the
water body reducing the depth of the photic zone. This reduces primary
production and decreases fish food. The NAS committee 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 sorptlon 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,
plant emergence and impedes soil aeration; the formation of films on plant leaves
which blocks sunlight and impedes photosynthesis and which may reduce the
marketability of some leafy crops like lettuce; and finally the adverse effect on
irrigation reservoir capacity, delivery canals and other distribution equipment
(NAS, 1974).
The criteria for freshwater fish and other aquatic life is essentially
that proposed by the N.A.S. and the Great Lakes Water Duality Board.
Ho?
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REFERENCES CITED:
Edberg, N. and Hofstan, B.V., 1973. Oxygen uptake of bottom sediment studied
in-situ and 1n the laboratory. Water Research, 7: 1285.
European Inland Fisheries Advisory Commission, 1965. Water quality criteria for
European freshwater fish, report on finely divided solids and Inland fisheries.
Int. Jour. A1r Water Poll., 9: 151.
Gammon, J.R., 1970. The effect of Inorganic sediment on stream biota. Environmental
Protection Agency. Water Poll. Cont. Res. Series, 18050 DWC 12/70, USGPO,.
Washington, D.C.
Mackenthun, K.M., 1973. Toward a cleaner aquatic environment. U.S.
Government Printing Office, Washington, D. C.
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria, 1972. U.S. Government Printing Office,
Washington, D. C.
Standard Methods for the Examination of Water and Wastewater, 1971,. 13th Edition.
APHA, Amer. Water Works Assn., Wtr. Poll. Cont. Fed.
Tarzwell, C.M. and A.R. Gaufin, 1953. Some Important biological effects of
pollution often disregarded in stream surveys, proceedings of the 8th Purdue
industrial waste conference. Reprinted 1n Biology of Water Pollution, 1967.
Dept. of Interior, Washington, D. C.
Ho*
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Tebo, L.B., Jr., 1955. Effects of sIHatlon, resulting from Improper logging,
on the bottom fauna of a small trout stream 1n the southern Appalachians.
The Progressive F1sh-Cultur1st, 17: 64.
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SULFIDE - HYDEQGEK SULFIDE
CRITERION :
?. ug/1 undlssoclated H-S for
fish and other aquatic life, fresh
and marine water
INTRODUCTION:
Hydrogen sulfide 1s a soluble, hlqhly poisonous, qaseous
compound having the characteristic odor
of rotten eqqs. It is detectable in air by humans at a
dilution of 0.002 ppm. It will dissolve in water at 400H
mg/1 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 deconnosition
of sewaqe, sludge beds, algae and other naturally
deposited organic material is a major source of hydrogen
sulfide.
When soluble sulfides are added to water thev react
with hydroqen ions to form HS""br t^S, the proportion of each
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depending on the pH. The toxicity of sulfides derives
primarily from HgS ra-ther than from the nydrosulfide (HS~)
sulfide (5*) ions. When hydrogen sulfide dissolves in water it-.
dissociates aocsordino to the reactions:
and
At pH 9 about 99 percent of the sulfide Is In the form
of HS~ at pH 7 the sulflde 1s equally divided hetv/een HS"*and
H2$ and at pf! 5 about 9.9 percent of the sulflde Is present
as H2S (HAS' 1974). The fact that H2S 1s oxidized In well-
aerated water by natural bioloqlcal systems to sulfates or
is biologically oxidized to elemental. ««lfwr h««i caused
Investigators to minimize the toxic effects of H2S on fish
and other aquatic life.
RATIONALE:
The degree of hazard exhibited by sulflde to aquatic
animal life 1s dependent on the temperature, r>H and
dissolved oxygen. At lower pH values a qreater proportion
1s In the form of the toxic undl ssociated H?S. In winter
when the pH 1s neutral or below or when dissolved
oxyqen levels are low but not lethal to fish, the hazard from
sulfides is exacerbated. F1sh exhibit a stronq avoidance
reaction to sulflde. Based on data from experlnents with
the stickleback, Jones (1964) hypothesized that 1f fish
-------
encounter a lethal concentration of sulflde there 1s a
reasonable chance they will be repelled by 1t before they
are harmed. This of course assumes that an escane route 1s
open.
Many past data on the toxldty of hydroqen sulflde to
fish and other aquatic Hfe have been based on extremely
short exposure periods. Consequently these early data have indicated that
concentrations between 0.3 and 0.4 my/1 permit fish to survive
(Van Horn 1958, Boon and Foil is 1967, Theede et al.. 1969). Recent
long-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/1 existed within 20 mm of the bottom of
sludge beds, and the levels of 0.1 to 0.02 mg/1 were common
within the first 20 mm of water above this layer. Walleye,
St1zpstedjon vltreum j eggs held 1n trays 1n this zone did
not hatch. Adelman and Smith (1970) reported that the
hatchabiHty of northern pike, Esox luclus ^ eaqs v/as
substantially reduced at 25 ug/1 HaS; at 47
uq/1 mortality was almost complete. Northern pike frv had
-------
96-hour LCso values that varied from 17 to 32 ug/1 at
normal oxygen levels of 6.0 mg/1. The highest concentration
of hydrogen sulfide that had no observable effect on eggs
and fry was 14 and 4 ug/1, respectively. Smith and Oseld
(1972), working on eggs, fry and juveniles of walleyes and
white suckers, Catostomus commersonl , and Smith (1971),
working on walleyes and fathead minnows, Plmephales promelas ,
found that safe levels varied from 2.9 ug/1 to 12 ug/1 with
eggs being the least sensitive and Juveniles being the most
sensitive In short-term tests. In 96-hour bloassays, fathead
minnows and goldfish, Carasslus auratus^ varied greatly In
tolerance to hydrogen sulflde with changes 1n temperature. They
were more tolerant at low temperatures (6 to 10°C). Holland,
et a_T_. (1960) reported that 1.0 mg/1 sulflde caused 100 percent
mortality 1n 72 hours with Pacific salmon.
On the basis of chronic tests evaluating growth and survival,
the safe HgS level for blueglll, Lepomls tracrochirus/ juveniles
and adults was 2 ug/1. Egg deposition 1n bluegills was reduced
after 46 days 1n 1.4 ug/1 H2S (Smith and Oseld, 1974). White
sucker eggs were hatched at 15 ug/1 , but juveniles showed
growth reductions at 1 ug/1. Safe levels for fathead minnows
were between 2 and 3 ug/1. Studies showed that safe lev'els for
Gammarus pseudolImnaeus and Hexagenla limbata were 2 and 15 ug/1,
respectively (Oseld and Smith, 1974a, 1974b). Some species
typical of normally stressed habitats. Asell us spp., were
much more resistant (Oseid and Smith, 1974c).
-------
Sulflde 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/1
undissociated H2S would not be hazardous to most fish and
other aquatic wildlife, but concentrations in excess of 2.0 ug/1
would constitute a long-term hazard.
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REFERENCES CITED:
Adelman, I.R. and L.L. Smith, 1970. Effect of hydrogen sulfide
on northern pike eggs and sac fry. Trans. Amer. Fish. Soc.,
99: 501.
Boon, C.W. and B.J. Follis, 1967. Effects of hydrogen sulfide
on channel catfish. Trans. Amer. Fish. Soc., 96: 31.
Colby, P.V. and L.L. Smith, 1967. Survival of walleye eggs and
fry on paper fiber sludge deposits in Rainey River, Minnesota.
Trans. Amer. Fish. Soc., 96: 278.
Holland, 6.A., et, a_l_., 1960. Toxic effects of organic and
inorganic pollutants on young salmon and trout. Wash. Dept.
Fish. Res., Bull. No. 5, 264 p.
Jones, J.R., 1964. Fish and river pollution. Butterworth and
Co., London, England.
National Academy of Sciences, National Academy of Engineering,
1974. Water quality criteria, 1972. U.S. Government Printing
Office, Washington, D. C.
-------
Oseld, D.M. and L.L. Smith, Jr, 1974a. Chronic ^oxldty of
hydrogen sulflde to Gammarus pseudolimnaeus. Trans. Amer.
F1sh.,Soc., 103 (1n press).
Oseld, D.M. and L.L. Smith, 1974b. Long-term effects of
hydrogen sulflde on Hexagenia limbata (Ephemeroptera).
Environmental Ecology (1n press).
Oseld, D.M. and L.L. Smith, 1974c. Factors Influencing acute
toxldty estimates of hydrogen sulflde to freshwater Inverte-
brates. Water Research 8 (1n press).
Smith, L.L., 1971. Influence of hydrogen sulflde on fish and
arthropods. Environmental Protection Agency, Project 18050
PCG, Washington, D. C.
Smith, L.L. and D.M. Oseld, 197/! Effects of hydrogen sulflde
on fish eggs and fry. Water Research,6: 711.
Smith, L.L.,, Jr., and D.M. Oseld, 1974. Effect of hydrogen
sulflde on development and survival of eight freshwater fish
species, Ijj: The early life history of fish. J.H.S. Blaxter
(ed.), p. 415-430. SpHnger-Verlag, New York, N.Y.
Theede, H.,et.aJ_., 1969. Studies on the resistance of marine
bottom Invertebratesto oxygen deficiencies and hydrogen
sulflde. Mar. B1ol... 2: 325.
Van Horn, W.M., 1958. The effect of pulp and paper mill wastes
on aquatic life. Proc. Ontario Indust. Waste Conf.» 5: 60.
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TAINTING SUBSTANCES
CRITERION;
Materials should not be present in concentrations that
individually or in examination 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 (MAS, 1974) lists a number of
wastewaters and chemical compounds that have been found to lower the palatabllity
of fish flesh. Implicated wastewaters included those from 2,4-D manufacturing
plants, kraft and neutral sulfite pulping processes, municipal wastewater treat-
ment 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/1 o-chlorophenol was reported to cause
tainting of fish flesh.
Shumway and Pal ensky (1973) determined estimated threshold concentrations for
HI?
-------
twenty-two organic compounds. The values ranged from 0.4 ug/1 (2,4-dichlorophenol)
to 95,000 ug/1 (formaldehyde). An additional twelve oonpounds were tested, seven of
which were not found to impair flavor at or near lethal levels.
Thomas (1973) reviewed the literature on tainting substances and listed
serious problems that have occurred; he detailed studies and methodology used in
the evaluation of the palatability of fishes in the Ohio River as affected by
various waste discharges. 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/1 of gasoline can impart off-
flavors to fish (Boyle, 1967).
418
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REFERENCES CITED;
Boyle, H.W., 1967. Taste/odor contamination of fish from the Ohio River.
Fed. Water Poll. Cont. Admin., Cincinnati, Ohio.
National Academy of Sciences, National Academy of Engineering, 1974.
Water quality criteria^ 1972. U.S. Government Printing Office, Washington, D.C.
Shumway, D.L. and J.R. Palensky, 1973. Impairment of the flavor of fish
by water pollutants. U.S. Environmental Protection Agency, EPA-R3-73-010,
U.S. Government Printing Office, Washington, D. C.
Thomas, N.A., 1973. Assessment of fish flesh tainting substances. In:
Biological methods for the assessment of water quality, J. Carlns and
K.L. Dickson (eds.) Amer. Society for Testing and Materials, Tech. Publ.
528, Philadelphia, Pennsylvania.
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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 time):
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/d(log f time | -a) - 2°C
(C°) V W 10 L (minr
where: l°9io = T°9aritnrn to base 10. (common logarithm)
a = intercept on the "y" or logarithmic axis of
the line fitted to experimental data and which
is available from Appendix II-C, NAS, 1974 for
some species.
b = slope of the line fitted to experimental data and
available from Appendix II-C, NAS, 1974 for some
species.
and
2. The second value is a limit on the weekly average temperature that:
a. in the cooler months (mid-October to mid-April in the north
and December to February in the south) will protect against
mortality of important species if the elevated plume temperature
is suddenly dropped to the ambient temperature, with the limit
being the acclimation temperature minus 2°C when the lower lethal
threshold temperature equals the ambient water temperature (in
some regions this limitation may also be applicable in summer).
or
In the warmer months (April through October in the north and
March through November in the south) is determined by adding
to the physiological optimum temperature (usually for growth)
a factor calculated as one-third of the difference between
the ultimate upper incipient lethal temperature and the
optimum temperature for the most sensitive important species
(and appropriate life state) that normally is found at that
location and time.
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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 require-
ments when they are applicable.
or
d. There Is a site-specific limit that is found necessary to
preserve normal species diversity or prevent appearance
of nuisance organisms.
.Marine Aquatic Life
In order to assure protection of the characteristic indigenous
marine community of a water body segment from adverse thermal effects:
a) the maximum acceptable increase in the weekly average
temperature due to artificial sources is 1° C (1.8° F)
during all seasons of the 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:
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 3.2.2° 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 reading ,.)
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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
Ihe uses of water by man in and out of its natural situs in the environment
are affected by its temperature. Offstream domestic uses and instream recreation
are both partially temperature dependent. Likewise, the 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 cooling is 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. r]o quote from
the FWPCA (1967), "Temperature, a catalyst, a depressant, an activator, a
restrictor, a stimulator, a controller, a killer, is one of the tiost important
and most Influential water quality characteristics to life in water."
RATIONALE
The suitability of water for total body immersion is greatly affected by
temperature. In temperate climates, dangers from exposure to loitf 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).
H2.2.
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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 over-
lying water and in bottom deposits which make§ 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 (19^4), Camp (1963), and Velz (1970).
Indicator enteric bacteria, and presumably enteric pathogens, are like-
wise affected by temperature. It has been shown that both total and fecal
coliform bacteria die away more rapidly in the environment with increasing
temperatures (Ballentine and Kittrell, 1968).
Temperature effects have been shown for water treatment processes. Lower
temperatures reduce the effectiveness of coagulation with alum and subsequent
rapid sand filtration. In one study, difficulty was especially .pronounced below
5°C (Hannah, et al., 1967). Decreased temperature also decreases the effect-
iveness of chlorination. Based on studies relating chlorine dosage to
temperature, and with a 30-minute contact time, dosages required for equivalent
disinfective effect increased by as much as a fadtor of 3 when temperatures
were decreased from 20° C to 10° C (Reid and Carlson, 197*0- Increased
temperature may increase the odor of water because of the increased volatility
of odor-causing compounds (Burnson, 1938). Odor problems associated with
plankton may also be aggravated.
-------
The effects of temperature on aquatic organisms have been the subject of
several comprehensive literature reviews (Brett, 1956; Pry, 1967; FWPCA, 196?;
Kinne, 1970) and annual literature reviews published by the Water Pollution
Control Federation (Coutant, 1968, 1969, 1970, 1971; Goutant and Goodyear,
1972; Goutant 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 (Cairns, 1956)". Likewise, changes from a
cold water 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, I960).
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
acclimation 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. The available data
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 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% of a sample of
organisms can survive is determined on the organism at the highest sustainable
acclimation temperature. The lowest temperature that 50% 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 which under other conditions of acclimation would not be
detrimental, may suffer a significant mortality due to 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 geographical distribution of species. Such
temperature changes are necessary to induce the reproductive cycles of
aquatic organisms and to regulate other life factors (Mount, 1969).
-------
Artifically Induced changes such as the return of cooling water or
the release of cool hypolimnetic waters from impoundments may alter indigenous
aquatic ecosystems (Coutant, 1972). Entrained organisms may be damaged by
temperature increases across cooling water condensers if the increase is
sufficiently great or the exposure period sufficiently long. Impingement upon
condenser screens, chlorination for slime control or other physical insults
damage aquatic life (Raney, 1969; Patrick, 1969 (b)). However, Patrick
(1969(a)) has shown that algae passing through condensers are not injured
if the temperature of the outflowing water does not exceed 34°£ to 34.5°C
In open waters elevated temperatures may affect periphyton, benthic
invertebrates, and fish in addition to causing shifts in algal predominance.
Trembley (I960) studied the Delaware River downstream from a power plant and
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 as currently defineable'requirements:
"1. Maximum sustained temperatures that are consistent with
maintaining desirable levels of productivity,
2. Maximum levels of metabolic acclimation to warm temperatures
that will permit return to ambient winter temperatures should
artificial sources of heat cease,
-------
3. Time dependent temperature limitations for survival of
brief exposures to temperature extremes, both upper and
lower,
4. Restricted temperature ranges for various states of reproduction
including (for fish) gametogenesis, spawning migration, release of
gametes, development of the embryo, commencement of independent
feeding (and other activities) by 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 o'f downstream aquatic life (in rivers)
where upstream diminution of a cold water 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 or 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 straight line
on a semi-logarithmic plot with time on the logarithmic scale and temperature
on the linear scale. (See Fig.p). In equation form this 50 percent
mortality relationship is:
-------
Iog10(time(mlnutes)) - a + b (Temperature(OG))
where: loglQ = logarithm to base 10 (common logarithm)
a = intercept on the "y" or logarithmic axis of
the line fitted to experimental data and w
is available from Appendix II-C, NAS, 1974
for some species.
b = slope of the line fitted to experimental data
which is available from Appendix II-C,. MAS., 1974
for some species.
To, provide a safety factor so that none or only a few organisms will perish,
it has been found experimentally that a 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
on size, life stage, and geographic source of the particular test species
(Appendix II-C, NAS, 1974)
Maximum temperatures for an extensive exposure (e.g., more than 1 week)
must be divided into those for warmer periods and winter. Other than
for reproduction, the most temperature-sensitive life function appears to be
growth (Coutant, 1972). Coutant (1972) has suggested that a satisfactory
estimate of a limiting maximum weekly mean temperature* is an average of the
optimum temperature for growth and the temperature for zero net growth.
* maximum weekly mean temperature - true mean temperature for a calendar
week which is ,a higher value than for any other week. Can be determined
from continuous measurements, hourly determinations, or some other non-
biased statistical method of analyzing temperature data.
-------
28
S
O
H
LU
O
LU
OC
LU
O.
LU
26
24
22
ACCLIMATION TEMPERATURE
24°
I . . . I .... I
70
100
1,000
. 10,000
TIME TO 50% MORTALITY-MINUTES
AFTER BRETT 1952
Figure 1. MEDIAN RESISTANCE TIMES TO HIGH TEMPERATURES AMONG YOUNG CHINOOK
(Oncorhynchus tshawytscha) ACCLIMATED TO TEMPERATURES INDICATED. LINE A-B
DENOTES RISING LETHAL THRESHOLD (incipient lethal temperatures) WITH INCREASING
ACCLIMATION TEMPERATURE. THIS RISE EVENTUALLY CEASES AT THE ULTIMATE
LETHAL THRESHOLD (ultimate upper incipient lethal temperature), LINE B-C.
(TAKEN FROM NAS, 1974)
129
-------
Because of the difficulty in determining the temperature of zero net growth,
essentially the same temperature can be derived by adding to the optimum
temperature (for growth or other physiological functions) a factor calculated as
1/3 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 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
JL1
extensive exposure for various fish species are presented in Table S. These
calculations use the above equations and data from ERL-Duluth, 1976.
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
2
for the "no stress" level from cold shocks (NAS, 1974). Figure, was
developed from available data in the literature (ERL-Duluth, 1976) and can
be used for estimating allowable winter temperature increases.
Coutant (1972) has reviewed the effects of temperature on aquatic life
reproduction and development. Reproductive events are noted as perhaps
the most thermally restricted of all life phases assuming other factors are at or
-------
TABLE
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 Growth5 Maximab
Atlantic Salmon 20 (68) 23 (73)
Bigmouth Buffalo
Black Crappie 27 (81)
Bluegill 32 (90) 35 (95)
Brook Trout 19 (66) 24 (75)
Carp
Channel Catfish 32 (90) 35 (95)
Coho Salmon 18 (64) 24 (75)
Emerald Shiner 30 (86)
Freshwater Drum - -
lake Herring (Cisco) 17 (63)c 25 (77)
Largemouth Bass 32 (90) 34 (93)
Northern Pike 28 (82) 30 (86)
Rainbow Trout 19 (66) 24 (75)
Sauger 25 (77)
Smallmouth Bass 29 (84)
Smallmouth Buffalo
Sockeye Salmon 18 (64) 22 (72)
Striped Bass
Threadfin Shad
White Bass
White Crappie 28 (82)
White Sucker 28 (82 )c
Yellow Perch 29. (84) -
a - Calculated according to the equation (using optimum temperature for growth)
maximum weekly average temperature for growth = optimum temperature
* 1/3 (ultimate incipient lethal temp. - optimum temperature
b - Based on temperature (DC) - 1/b (log-jg tl-|lie(min.) ~a) " 2°c» acclimation
at the maximum weekly average temperature for summer growth, and data
in Appendix II-C of Water Quality Criteria, 1972 (NAS, 1973).
c - Based on data for larvae (ERL-Duluth, 1976).
-------
C0
CM
432
-------
near optiinum levels. Unnatural short-term temperature fluctuations appear
to cause reduced reproduction of fish and invertebrates. There are in-
adequate 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 anadroRjous species or other migrants who pass frcm the
warmed area to a normally chilled, unproductive area. Reported temperature
data on maxinum temperatures for spawning and embryo survival have been
1Z
summarized in T^ble (from ERL-Di/luth 1976).
Although the limiting effects of thermal addition to estuarine and
marine waters are not as conspicuous in the fall, winter and spring as during the
summer season of maximum heat stress, nonetheless crucial thermal limitations
do exist. Hence, it is important that the thermal additions to the receiving
waters be minimized during all seasons of the year. Size of harvestable
stocks of commercial fish and shellfish, particularly near geographic limits
of the fishery, appear to be markedly influenced by slight changes in the
long-term temperature regime (Dow, 1973).
Jefferies and Johnson (1974) studied the relationship between temperature
and annual variation in 7-year catch data for winter flounder, Pseudopleuronectes
americanus. in Narragansett Bay, Rhode Island. A 78 percent decrease in annual
catch correlated closely with a.0.5°C increase in the average temperature over
the 30-month period between spawning and recruitment into
-------
TABLE
Summary of Reported Values for
Maximum Weekly Average Temperature for Spawning and Short-Term
Maxima for Bribryo Survival During the Spawning Season
(Centigrade and Fahrenheit)
Embrvo
Species
Atlantic Salmon
Bigtr.outh Buffalo
Black Grapple
Bluegill
Brook Tr^L
Carp
Channel Ct^rt^l:
Cohe Salmon
Herald Shiner
Freshwater Drum
Lake Herring (Cisco)
Largemouth Bass
Northern Pike
Rainbow Trout
Sauger
StnalliTioxith Bass
Stoallmouth Buffalo
Sockeye Salmon
Striped Bass
Threadf in Shad
White Bass
White Grapple
White Sucker
Yellow Perch
SpawnlngQ-
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)
Survival"
7
27
—
34
13
33
29
13
28
26
8
27
19
13
21
-
21
13
24
34
26
23
20
20
(45)
(8l)c
(93)
(55)
(91)
(84)
(55)
(82)c
(79)
(46)
(81)
(66)
(55)
(70)
(70)
(55)
(75)
(93)
(79)
(73)
&#
(68)
a - the optimum or mean of the range of spawning temperatures reported
for the species (ERL-Duluth, 1976).
b - the upper temperature for successful incubation and hatching reported
for the species (ERL-Duluth, 1976).
c - upper temperature for spawning
-------
the fishery. Sissenwine's 1974 model predicts a 68 percent reduction of
recruitment in yellowtail flounder, Limanda ferruglnea, with a 1° C long-
term elevation in southern New England waters.
Community balance can be influenced strongly by such temperature de-
pendent 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 inter-species relation-
ships. A 50 percent reduction in the soft-shell 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 four 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 wintpr cnld kill of the green crab
and permitted a major expansion of its population, with increased predation
i**» I9fe«
of the soft-shell clam resulting (Slude, ; Welch, - ).
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 there are some species which
-------
spawn only during the fall (herring), 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 £. balanoides . the edible blue mussel Mytilus edulis . sea urchin ,
Strongyiocentrotus drobachiensi s . sculpin, and the winter flounder,
b0pgjj£_
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 consequencei fishes usually are attracted to heated water during the
fall, winter, and spring. Avoidance will occur as water 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 possi-
bility of cold shock and death of plume-entrained fish due to 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 due to insufficient food (Massengill, 1973). A weight
loss of approximately 10% for each 1°C rise in water temperature has been observed
in fish when food is absent (Phillipset_al., I960) There may also be indirect
adverse effects on the indigenous community due to increased predation
pressure if thermal addition leads to a concentration of fish which are
dependent on this community for their food.
Pish 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 of anadromous fish into rivers (Leggett and Whitney, 1972).
In the autumn, a number of juvenile marine fishes and shrimp are dependent
on a drop in temperature to trigger their migration from estuarine nursery
grounds for oceanic dispersal or southward migration (Lund and Maltezos,
1970; Talbot, 1966).
Thermal discharges should not alter diurnal and tidal temperature
variations normally experienced by marine communities. Laboratory studies
show thermal tolerance to be enhanced when animals are maintained under a
diurnally fluctuating temperature regime rather than at a constant temperature
19*75
(Costlow and Bookhout, 1971; Purch, 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 shdVild be established to protect the various
marine communities within each biogeographic region. During the summer,
naturally elevated temperatures may be ofaufficient magnitude to cause
death or emigration (Chin, 1961; Glynn/'Vaughn, 1918) 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, 197(3; Zieman, 1970). The general
ceilings set forth here are derived from studies delineating limiting tem-
peratures for the more thermally sensitive species or conmunities 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 utilized to develop thermal limits which
specifically consider communities relevant to given water bodies.
-------
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-------
Trembley, F.J., I960. Research project on effects of condenser discharge
water on aquatic life, progress report, 1956-1959. The Institute of
Research, Lehigh University.
Vaughan, T.W., 1918. The temperature of the Florida coral-reef tract.
Publs. Carnegie Inst. Wash., 213: 319-
Velz, C.J., 1970. Applied stream sanitation. Wiley-Interscience, New York.
Welch, W.R., 1968. Changes in abundance of the green crab, Carcinus rnaenas
(I/.), in relation to recent temperature changes. Pish. Bull. 67, 2: 337-
Zieman, J.C., Jr., 1970. The effects of a thermal effluent stress on the
sea grasses and macroalgae in the vicinity of Turkey Point, Biscayne Bay,
Florida. Ph.D. Dissertation, Univ. Miami, Coral Gables, Fla., 129 p.
-------
Appendix
Estugyine and Cpastal^Biota
Recommended marine thermal criteria are based on scientific
evaluation of available data. Representative thermal effects
data are summarized here for an array of ecologically diverse
marine organisms, grouped by biotic region. Since the summer
temperature regime can provide "worst case" thermal conditions,
studies dealing with warm-acclimated organisms are cited
primarily. Findings of sublethal effects studies are listed
also. Twenty-four-hour TLm (median tolerance limit) data have
been adjusted by subtracting 2.2°C to estimate the upper thermal
protection limit for the life history stage in question
(Mihursky, 1969) . Recognized biological variables such as recent-
environmental history, nutritional state, size, sex, and age have
been considered for all thermal effects investigations.
Likewise, contrasting methods of study were considered.
Normally, thermal effects data derived in one biotic region
should not be applied to another. Latitudinal ly separated
populations of widely distributed species may exhibit significant
genetic variability and usually have experienced different,
recent environmental histories. Boundaries for regional ceilings
are demarcated by biogeographic provinces. Species composition
of the marine system, and most important:, responses to elevated
-------
temperature, are generally similar within a region. Boundaries
of a biotic province are characterized by significant thermal
discontinuities. Boundary areas are maintained during summer or
winter due to combined forces of current, wind, and coastal
geomorphology. On the east coast, Cape Canaveral, Pla.; Cape
Hatteras, North Carolina; and Cape Cod, Massachusetts, represent
these boundaries. On the west coast. Point Conception in
southern California marks the limit of warm and cold temperate
zones.
Atlantic Coast: This region extends from Cape
Cod, Mass. , to the Gulf of Maine. Insufficient data are
available for setting regional temperature limits. Upper limits
should be determined on a case-by-case basis using best available
data for the site and its environs.
In the boreal region, maintenance of a general temperature
regime resembling natural conditions is particularly important
during winter months. Some boreal species require periods of
uninterrupted low water temperatures to fulfill environmental
requirements for successful maturation of sexual products,
spawning, and subsequent egg and larval survival. Winter
flounder, Pseudopleuronectes americanus, have an upper limit for
spawning of 5.5°C (Bigelow and Schroeder, 1953) . Spawning occurs
during the winter.
-------
Ten degrees centigrade is the upper thermal limit for
Atlantic salmon, Salmg salar, smolt migration to the sea, which
normally occurs in June. Twelve degrees centigrade inhibits
maturation of sex products (DeCola, 1970). Development of winter
flounder, gsgudop^euronec tes americanus, eggs to hatching is
reduced 50 percent at 13°C (Rogers, in press). Blood worm,
gj.ycera americapa, spawning is induced when temperatures reach
13°C (Creaser, 1973). Fifteen degrees centigrade is the upper
limit for spawning Atlantic herring, clupea harengus, (Hela and
wz>
Laevastu, and of an araphipod, Psammonx nobj,j.4g* (Scott,
1975) . In Atlantic herring, there is above normal incidence of a
protozoan disease at 15°C (Sinderman, 1965) and at 16°c, there is a prevalence
of erythrocyte degeneration (Sherburne, 1973). Field mortality of
yellowtail flounder larvae, Lj.manda ferr.uain.ea, was observed at 17.8°c
(Colton, 1959). The protection limit for yearling Atlantic
herring, Clugea harengus, (48-hr. TLm - 2.2°C) is 19.0°C (Brawn,
1960). At 21°C, embryonic development ceases in the amphipod,
gammarus duebeni, (Steele and Steele, 1969). Above 21.2°c,
spores are killed and growth is reduced in the macroalga,
ChpUdrus crispus, which is commercially harvested as Irish moss
(Prince and Kingsbury, 1973).
.£old r.£2E££§te j|onex Atlantic Coast: Temperature ceilings
are particularly critical in the southern portion of this region
."south shore of Long Island to Cape Hatteras, N.C.) where
nclosed sounds and large coastal-plain bays and rivers are
HH9
-------
prevalent. Maximum temperatures should not exceed 30.6°C. Were
temperatures of 30°C to persist for over 4 to 6 hours,
appreciable stress or direct mortality would occur among juvenile
winter flounder, Pseudopleuronectes americanus; striped mullet
larvae, Mugil ceghaJLus; Atlantic silverside eggs and adults,
£§Bi<|i.a menidia; adult northern puffer, SjDhaeroides SiiSliiat us ;
adult blue mussel, MXtiius edul^is; and adult soft shell clam, Mya
ajrenaria. Specif ic critical temperatures for these species are
detailed in Table The adult protection limit (TLm - 2.2°C)
is 28.8°C for sand shrimp, Crancjon septemspinosa. and 30.8°C for
opossum shrimp, Neoisv.sis amer ican\^s . Both are important food
organisms for fish (Mihursky and Kennedy, 1967) . Respiration
rate is depressed above 30°C in the mole crab, (Edwards and
Irving, 1943). At 31.5°C, there is 67 percent mortality in coot
clams when exposed for 6 hours (Kennedy, et al. , 1974).
A true daily mean limit of 27.8°C approximates the upper limit for larval
growth of the coot clam (27.5°C; Calabrese, 1969). Between 28<>c
and 30°C juvenile amphipods, Corophium jLnjgidi os\un , leave their
tubes and thereby lose natual protection from predation
(Gonzalez, 197A . such elevated temperatures may also have
subtle sublethal effects, such as reducing feeding and growth.
In the quahaug, Mercenaria roe r£§na ri a , growth is optimum at 20°C
(Ansell, 1968) . Growth is inhibited above 24°C in a rock weed,
§U£!» (South and Hill, 1970). Prolonged
-------
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locomotion is markedly reduced at 22°C in the Jonah crab, Cancer
l&T.f&IiS' at 28°c in ££&.£§£ irroratus (Jeffries, 1967). An
oyster pathogen, Democystidium marjnum, readily proliferates
above 25°C (Andrews, 1965).
High temperature usually will elicit avoidance response in
fishes. Avoidance is triggered at 29°C in Atlantic menhaden,
grgyooictj.a tyrannus, and at 26.5°C in sea trout, Cyjioscion
rggalis, (Meldrin and Gift, 1971). Breakdown of the avoidance
response in striped bass, flip rone saxatilis. occurs at 30°C (Gift
and Westman, 1971). Maximum reported temperature for capture of
spotted hake, Prophycig regius, is 2*,8°C in the Chesapeake Bay
(Barans, 1972),
North of Long Island, a 1.0°C rise above summer ambient pro-
vides reasonable protection. For example, maximum short-term
temperatures in Narragansett Bay, Rhode Island, usually would not
exceed 23.1°C in August, judging from 15-year mean temperature
data for Fox Island . Larval Atlantic silverside, juvenile
winter flounder, and blue mussel should be protected by that
*
thermal limitation. The thermal protection limit (TLm - 2.2°C)
for juvenile winter flounder is 26.9°C (Gift and Westman,
H52.
-------
Repeated exposures to 25°C would stress the blue mussel, >fytilus v
edulis . by causing cessation of feeding (Gonzalez, 1972). Diurnal
summer. maxima exceeding 22°C can alter normal metabolic rates in
embryonic tautog, Tautqga_ onitis^ (Laurence, 1973) and cause feeding
problems for adult winter flounder (Olla, 1969) and the sand-collar
snail, Polinices duplicata , (Hanks, 1953).
Optimum for summer development of the rock crabf Cancer
irroratus, and Jonah crabf £. boreal is f larvae is 20°C; at 25°C,
mortality precludes completion of larval development (Sastry and
Vargo, in press). Between 15 and 20°C, activity of the amphipodf
Gammarus oceanicusj is much reduced (Halcrox* and Boyd, 1967). Ini-
tiation of spawning is often cued by temperature. Blue mussel spawning
occurs when spring temperatures reach 12°C (Engle and Loosanoff, 1944).
A minimum of 10°C is required for their embryonic development
CBrenko and Calabrese, 1969) and spawning occurs at 15°C. Peak
spawning runs of American shad, Alosa sapidissima , into rivers occurs
at 19.5°C (15 year average, Connecticut River); downstream migration
of juveniles occurs as temperature falls below 15.5°C (Leggett and
Whitney, 1972). Menhaden migrate at 10°C (Bigelow and Schroeder,
1953); striped bassp Morone saxatilis, migrate into or leave rivers
at 6 to 7.5°C (Merriman, 1941). In the fall and winter, fishes con-
gregate in discharge plumes which exceed these temperatures. These
fishes exhibit increased incidence of disease and a general loss of
physiological condition (Mihursky, et al; 1970).
£oneA Atlantic and Gulf £oasts: This region
:tends from Cape Hatteras, N.C., to Cape Canaveral, Florida, and
-------
on the Gulf Coast from Tampa1, Florida, to Mexico. The
recommended regional ceiling is a short-term maximum of 32.2°C .
Prolonged exposures to temperatures near this level would
adversely affect portions of the biota. At 33°C, bay anchovy,
&Q£&°£ m^tchilli, embryonic development is reduced to 50 percent
of optimum (Rebel, 1973). The upper limit for growth of juvenile
white shrimp, Penaeus setifejrus, is 32«5°C (Zein-Eldin and
Griffith, 1969). A decline in field abundance of brown shrimp,
Penaeus az£ecus, at temperatures above 30 °C was reported by Chin
(1961).
Protection limits {50 percent of optimal survival) of two
sardines, Harengu3,a iaguana and Harencjula pensacolae. for
development of the yolk sac larval stage are 31.<*°C and 32.2°C,
respectively {Rebel, 1973; Sakensa, et al., 1972). Larval
pinfish, Lagodon rhomboides, and spot, Leigstoinus xarjthurus,
exhibit a breakdown in avoidance response mechanisms at 31.0°C
and 31. 1°C, respectively (Boss, D.E., etaL., 1974).
The protection limit (TLm - 2.2°C) for young-of-the-year Atlantic
menhaden is 30.8°C (Lewis and Hettler, 1968). Upper limit for
adult growth of the quahaug, Mercenarj.a me^ceparia, is 31°C
(Ansell, 1968).
Daily mean temperatures continually exceeding 29°C would
result in mortality of striped mullet eggs, MugjLl. cephalus.
-------
Their 96-hour TLm is 26.U°C (Courtenay and Roberts, 1973). Egg
and yolk sac larval survival of sea bream, Archosargus rhomboi-
<23li§' is reduced to 50 percent of optimal at 29.1°C. For
yellowfin menhaden, Breyoortia smithi, exposure to 29.8°C reduced
survival of egg and yolk sac larvae to 50 percent of optimal
(Rebel, 1973). Sublethal but potentially damaging ecological
effects could occur at levels well below 29°C. For example, the
upper limit for optimal growth of post-larval brown shrimp,
££Q££gJiS a2t§£ili» is 27.5°C (Zein-Eldin and Alrich, 1965).
Developing embryos and fry of striped bass cannot tolerate 26.7°c
in fresh water (Shannon, 1969). This report may also apply to
fry in waters at ,the head of estuaries. This species spawns in
early spring.' Elevation of winter temperatures above 20°C in St.
Johns River, Florida, could interfere with upstream migration of
American shad, AJosa sapidissima, (Leggett and Whitney, 1972).
Subr Tropical Regions; Ceilings for sub-tropical regions such as south
Florida (Cape Canaveral and Tampa southward), and Hawaii are an in-
stantaneous maximum 32.2°C and a true daily mean not exceeding 29.4OC.
Ceilings for true tropical sites should he developed from studies of
indigenous populations of relevant communities. Physiological variation
In thermal adaptations and tolerances have been reported for coral be-
tween sub-tropical (Hawaii, 19-22° Lat.) and true tropical sites
(Eniwetok Atol, Marshall Islands, 11° Lat.) (Jokiel, £t al., in press).
Much of the following thermal effects data represent southern
Florida or Hawaiian biota. -A review by Zieman and Wood (1975)
•uggests that the thermal optimum is 26-28°C for tropical marine
systems, with chronic exposure to temperatures between 28°C and
30°C causing heat stress. Death of the biota is readily discernible
totveen 30°C and 32°C. Hayer (1914) recognized that nearshore
-------
tropical marine biota normally live at temperatures only a few
degrees below their upper lethal limit. A study of elevated
temperature effects on the benthic community in Biscayne Bay,
Florida, resulted in the following data (Boessler, 1974):
Temperature for High
Temperature for 50 Percent
Species Exclusion I°C^
Molluscs
Echinoderms
Coelenterates
Porifera
26.7
27.2
25.9
2U.O
31.«
31.8
29.5
31.2
Other" thermal data for tropical biota include a 25.4-27.8°C
optimum for fouling community larval settlement (Roessler, 1974);
25°C optimum for larval development of j'oj.yonYx gibbesi, a
commensal crab (Gore, 1968); 27°C for growth and gonad develop-
ment in sea urchins, SatgsbiQJJS variegatus, and for growth -in a
snail, Cantharus £iS£tus, (Albertson, 1973); 27 to 28°C optimum
for larval development of pink shrimp, Peoaeus duorarum.
(Thorhaug, et aj.., 1971a); and 30°C optium for turtle grass,
'yhalassj.a testud^num. productivity (Zieman, 1970). Kuthalingham
(1959) studied thermal tolerance of newly hatched larvae of ten
tropical marine fishes in the laboratory. When held at a series
of constant temperatures for 12 hours immediately following
-------
hatch, optimal survival for all species fell between 28°c to
30°C, but their tolerance limit ranged from 30°C to 32°C.
Thermal stress of the fouling community is seen in a 50 per-
cent reduced settlement rate of larvae at 28°C {Roessler, 1974).
Fifty percent reduction in gonadal volume of the sea urchin,
i2£££liilius varlegatus/ occurs at 29.9°C (Thorhaug, et al., 1971b).
Irreversible plasmyolysis of the macroalga, Valon^a ventrj-cosa,
at 29.9°C and of Valonia macroghysa at temperatures above 29.7°C
has been reported. Survival of developing embryos to the yolk
sac larval stage was reduced to 50 percent of optimal at 29,1°C
among sea bream, Archosargus rhomboidalis. At 29.8°c, yellowfin
menhaden, Breygortia smithi, and at 31.4°c scaled sardines,
SS££B3Mi^ •jaguana, suffer similar mortalities during early
development (Rebel, 197.3) . Temperatures in excess of 31°c to
33°C can interfere with embryonic development in six species
of mangrove-associated nematodes, even though adults can tolerate
an additional 2°C to 7°C {Hopper, et al., 1973). Upper limit for
larval (naupliar) metamorphosis in pink shrimp, Penaeus duorarum,
is 31.5°C {Thorhaug, et al., 1971b). Upper lethal temperatures
include 31.5°C for five species of Valonia {Thorhaug, 1970);
death in 3 to 8 hours for five Hawaiian corals at 31-32°C
(Edmondson, 1928; Jokiel and coles, 1974); 32°C Tim {95-hour) for
the sea sguirt, Ascidia riigra, and sea urchin, Lytgchinus
yariegatis, {Chesher, 1971) . Average daily temperatures near 31°C
-------
for three to ten days result in decreased growth in seagrass,
Thai,Cassia testudinum and red macroalgae, Laurencia fioitei.
Between 32°C and 33°C, health and abundance of these species
declines markedly (Thorhaug, 1971, 1973). Replacement of
seagrass is slow, especially if rhizomes are damaged due to
excessive consumption of stored starch during heat stress
(Zieman, 1970). Recovery of ThalJLassia beds may take decades
1*15
(Zieman and Wood, *»»,««'•—'«"•«*) .
Pacific Coast: Fewer thermal effects studies have been
conducted on West Coast species. However, the concept of sea-
sonal restrictions for temperature elevations above ambient are
well supported in several East Coast provinces and is deemed
applicable to the West Coast as a general biological principle.
Data are not sufficient to develop general regional ceilings.
These must be determined on a case-by-case basis until general
principles emerge.
The Pacific Coast consists of two distinct biogeographical
regions; the cold temperate province ranges north from Point
Conception, California and the warm temperate region from Point
Conception southward. Published data should provide a general
indication of possible adverse effects of excessive thermal
discharge on indigenous species.
-------
S £2is! 2!£ffl£§£2^§ 1°JQ£* Some winter and spring spawning
temperature ranges include 3°C to 6°C for Pacific herding, Clugea
Eailasi, (McCauley and Hancock, 1971); 7°C to 8°C for English
sole, Paroghr^s vetuius, (Alderdice and Forrester, 1968); 13°C for
May and June spawning of razor clams, Siligua ga^uj.a, {McCauley
and Hancock, 1971); and 12°C to 1U°C for native little neck
clams, Protothaca staminea, (Schink and Woelke, 1973). Optimal
growth occurs at 10°C in the small filamentous red algae,
Antithamnion spp., (West, 1968), and 12°C to 16^6 is optimal for
growth and reproduction of various red and brown algae, including
kelp, MacrogYgtig 2Y£JLf£r3» (Druehl and Hisiao, 1969) , Twelve to
16°C favors sea grasses, Zostera marina and Phylj.ospadix scouj.eri
(McRoy, 1970) . Spawning migration of striped bass, Mororje
saxatilis, occurs at 15°C to 18°C (Albrecht, 1964); in American
shad, Aj.osa sagidissima, spawning runs occur at 16.0-19.5°C
{Leggett & Whitney,1972). At Vancouver Island, B.C.,
distribution of a kelp, Ija.njin.aria. groenlandica ., is temperature
influenced . The long stipe form is not found above 13°C; the
short stipe form does not occur above 17°c . Tn the laboratory,
elevation of temperature to 13°C produces abnormal sporophytes
(Druehl, 1967) . Dungeness crab, Cancer magjster, larval
development is optimal at 10 and 13.9°C, survival is reduced at
17.8°C, with no survival to megalops at 21.7°C (Reed, 1969). The
Upper thermal limit for razor clam embryonic and larval develop-
ment is 17°C (McCauley and Hancock, 1971). Upper growth limit
-------
for small filamentous red algae, Antithamnion spp, is 13°C (W»st,
1968). King salmon migration into the San Joaquin River may be
delayed by estuarine temperatures in excess of 17.8°C (Dunham,
1968) .
The sea grass, Phy.llosp.adix scouleri, begins to die off
(McRoy, 1970), and the pea pod borer, Potula fulcata, ceases to
develop at 20°C (Fox and Corcoran, 1957) . Twenty degrees
centigrade also is the upper limit for embryonic and larval
development of the summer-spawning horse clam, Tresus nuttaj-li.,
and native little neck clam, Protothaca .sjtaminea, (Schink and
Woelke, 1973). Upper incipient lethal temperature tor the mysid
shrimp, Neomysis intermedia, is 21.7°c
-------
Gonadal recrudescence will not occur at 24°C or above, regardless
of photoperiod (DeVlaming, 1972). The 36-hour TLm for red
abalone adults is 23°C when acclimated to 15°C; for the embryos,
26°C, when exposed for 30 hours (Ebert, 1974). Sea urchin,
g^irongvlocentrotus gurguratus, upper tolerance limit is 23.5°C
for adults (Gonor, 1968); 25°C is lethal to embryos and renders
adults limp and unresponsive after H hours (Farmanfarmaian and
Giese, 1963) .
£ac,ific Warm Temperate 3£jje: The thermal threshold for
spawning in Pacific sardine, Sardingps caerujlea, is 13°C (Marr,
1962). Reports of temperature optima for spawning include 15°C
in a ctenophore, P^eurobranchia bachei, (Hirota, 1973); 16°C in
the spring spawning wooly sculpin, C^inpcottus ariaj.is, (Graham,
1970); 17.5°C for northern anchovy, Enqraulis mgrdax; 19°C for
opaleye fish, Gire^lg Qigricajjs, (Nprris, 1963) . Larval survival
is best at lg to 18°C in white abalone, galio.t^s sorenseni,
(Leightpn, 1972) .
Limiting effects of temperature include scarcity of the kelp
isopod in the beds above 17.8°C (Jones, 1971). Upper limit for
growth in Pleurobranchia bac.hei is 17°c; 20°C is the upper
tolerance limit for the adult ctenophore (Hirota, 1973). Twenty
degrees centigrade also causes limited survival in recently
settled juvenile white abalone (Leighton, 1972). Limiting
Hfel
-------
effects for wooly sculpin include the upper limit.of optimal
growth at 21°C; at 22°C, a 50 percent reduction in the successful
development of eggs; at 2H°C, the upper limit tor embryonic
development is reached (Hubbs, 1966) . Sea Urchins,
S^rongylocentrotus sp_. are weakened or killed at 2U°C to 25°C
(Leighton, 1971) . /\t 25°C, partial osmoregulatory failure occurs
in staghorn sculpin, I?2£t.2£2t!^§ 5SS.iiJSII» at 37,6 o/oo fMorris,
1960). A maximum temperature of occurrence of 25°C is reported
for top smelt, AtherJ.noj3S affinis, by Doudoroff (1945) and
northern anchovy, I?ngraul.is mordax, (Baxter, 1967) . For
,topsmelt, the upper limit at which larvae hatch is 26.8°C (Hubbs,
1965) .
Natural summer temperatures are stressful to beds of giant
kelp, MacrocYStis gyjrifera, in southern California. This
precludes any summer thermal discharge in the vicinity of these beds.
Deterioration of surface blades is evident from late June onward,
due in part to reduced photosynthesis (Clendenning, 1971) .
Several weeks» exposure to 18.9°C is harmful to the beds (Jones,
1971), while temperatures over 20°C result in pronounced loss of
kelp (North, 1964) . Brandt (1923) reported that there was scma 60 percent
reduction of kelp harvest when the average temperature was
20.65°C and that a bacterial disease, black rot, thriven on kelp
at 18-20°C. One-day exposure to 22°C is quite harmful to
cultured gametophytes of giant kelp (North, 1972) .
-------
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-------
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-------
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ZINC
CRITERIA:
5,000 ug/1 for domestic water supplies (welfare);
For freshwater aquatic life, 0.01 of the 96-hour
LCgQ as determined through bioassay using a
sensitive resident species.
INTRODUCTION:
Zinc usually is found in nature as the sulfide; it is often
associated with sulfides of other metals, especially lead, copper
cadmium, and iron. Most other zinc minerals probably have been -, >rme
as oxidation products of the sulfide; they represent only minor soura
of zinc. Nearly 3,000,000 short tons of recoverable zinc per year are
mined in the world; about 500,000 tons of this come from the United
States.
Zinc (as metal) is used in galvanizing, i.e., coating (hot dipping
of various iron and steel surfaces with a thin layer of zinc to retard
corrosion of the coated metal. In contact with iron, zinc is oxidized
preferentially, thus protecting the iron. The second most important use
of zinc, reaching major proportions in the last quarter century, is in
the preparation of alloys for dye casting. Zinc is used also in brass
and bronze alloys, slush castings (in the rolled or extruded state),
in the production of zinc oxide and other chemical products, and in
photoengraving and printing plates.
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Kopp and Kroner (1967) report that in 1,207 positive tests for zinc on
samples from U.S. waterways, the maximum observed value was 1,183 ug/1
(Cuyahoga River at Cleveland, Ohio) and the mean was 64 ug/1. Dissolved zinc
was measured in over 76 percent of all water samples tested. The highest mean
zinc value, 205 ug/1, was found in the Lake Erie Basin, whereas the lowest
mean zinc value, 16 ug/1, was observed in the California Basin. In seawater,
zinc is found at a maximum concentration of about 10 ug/1.
RATIONALE:
Zinc is an essential and beneficial element in human metabolism
(Vallee, 1957). The daily requirement of preschool-aged children is
0.3 mg Zn/kg body weight. The daily adult human intake averages 0 to
15 mg/ zinc; deficiency in children leads to growth retardation. Coro-
.munit.y water supplies have contained 11 to 27 mg/1 without narmful effects
(Anderson, e_t a]_., 1934; Bartow and Weigle, 1932). However, in tests
performed by a taste panel, 5 percent of the observers were able to
distinguish between water containing 4 mg/1 zinc as ZnS04, which had
a bitter or astringent taste, and water containing no zinc salts (Cohen,
e_t al_., 1960). Because zinc in water produces undesirable aesthetic
effects, the concentration of zinc in domestic water supplies should
be below 5 mg/1 (5,000 ug/1).
The toxicity of zinc compounds to aquatic animals is modified by
several environmental factors, particularly hardness, dissolved oxygen,
and temperature. Skidmore (1964), in undertaking a review of the
literature on the toxicity of zinc to fish, reported that salts of the
-------
alkaline-earth metals are antagonistic to the action of zinc salts,
and salts of certain heavy metals are synergistic in soft water. Both
an increase in temperature and a reduction in dissolved oxygen increase
the toxicity of zinc. Toxic concentrations of zinc compounds cause
adverse changes in the morphology and physiology of fish. Acutely toxic
concentrations induce cellular breakdown of the gills, and possibly
the clogging of the gills with mucous. Chronically toxic concentrations
of zinc compounds, in contrast, cause general enfeeblement and widespread
histological changes to many organs, but not to gills. Growth and
maturation are retarded.
Using dilution water with calcium of 1.7 mg/1 and magnesium of 1.0
mg/1, Affleck (1952) found a 54 percent mortality of rainbow trout fry
in 28 days in a zinc concentration of 10 ug/1. Pickering and Henderson
(1966) determined the 96-hour LCgQ of zinc for fathead minnows, Pimephjles
promelas, and bluegills, Lepomis macrochi rus, using static test conditions.
For fathead minnows in soft water (20 mg/1 CaC03) the LC50 was 870 ug/1,
and in hard water (360 mg/1 CaCOj) it was 33,000 ug/1. Bluegills were
more resistant in both waters. Similarly, the lethal threshold concen-
tration was 3 or 4 times as high for coarse fish as for trout, Salvelinus
fontinalis, (Ball, 1967).
The 24-hour LC$Q of zinc for rainbow trout, Salmo gairdneri. was
reduced only 20 percent when the fish were forced to swim at 85 percent
pf their maximum sustained swimming speed (Herbert and Shurben, 1964).
The maximum effect of a reduction in dissolved oxygen from 6 to 7 mg/1 to
2 mg/1 on the acute toxicity of zinc was a 50-percent increase (Lloyd,
1961; Cairns and Scheier, 1958; Pickering, 1968).
-------
The Atlantic salmon, Salmo salar, was tested in a 168-hour
continuous-flow bioassay at 17° C in water with a total hardness of
14 mg/1 CaCO.t. The incipient lethal level, the level beyond which
the organism can no longer survive, was 420 ug/1 of zinc (Sprague
and Ramsay, 1965).
Brungs (1969) found that in water with a total hardness of 200 mg/1
CaC03, 180 ug/1 zinc caused an 83 percent reduction in eggs produced by the
fathead minnow, Pimephales promelas, in chronic tests. The tests lasted
10 months and the control test water contained 30 ug/1 zinc. The
96-hou'r continuous-flow TLm was determined to be 9,200 ug/1 zinc.
A number of short term fish toxiclty data are detailed In Table V.
When referring to this table, the reader should consider the species
tested, pH, alkalinity, and hardness 4f alkalinity is not given (in most
natural waters alkalinity parallels hardness). In general, the salmonidfle
are most sensitive to elemental zinc in soft water; the rainbow trout
Salmo gairdneri is the most sensitive in hard waters. The influence of
the pH on solubility of zinc complexes, and the resulting toxicity of
zinc, is clearly shown in the data presented on the fathead minnow,
Pimephales pjrpmelas. The influence of pH and other factors on the
solubility and form of the zinc preclude the recommendation of a
freshwater or marine criteria based on acute toxicities alone.
-------
Wurtz (1962) performed bioassays on young pond snails, Physa
heterostropha. in waters with a total hardness of 100 mg/1 and
20 mg/1 CaC03. In water with a tejnperature of 51° F, the soft water tests
resulted in a 96-hour LC50 of 303 ug/1 zinc, whereas the 96-hour LC50
in hard water was 434 ug/1 zinc. The LC50 of a zinc sulfate solution
in dilution water with a total hardness of 44 mg/1 CaCOj for a
10-day test to a mayfly, Ephemerella subvaria. was 16,000 ug/1 (Warnick
and Bell, 1969). In such tests with several heavy metals, the immature
insects seem to be less sensitive than many fishes that have been tested.
The 48-hour LCRn for Daphnia magna in soft water with a hardness of
45 niK/1 CaGOj and an alkalinity of 42 mg/1 has been found to be
100 ug/1; in 70 ug/1 zinc, there was a 16 percent reproductive impairment
in a 3-week chronic test (Biesinger and Christensen, 1972).
Toxicities of zinc in nutrient solutions have been demonstrated for
a number of plants. Hewitt (1948) found that zinc at 16 to 20 mg/1
produced iron deficiencies in sugar beets. Hunter and Vergnano (1953)
found toxicity to oats at 25 mg/1. Millikan (1947) found that 2.5 mg/1
produced iron deficiency in oats. Early (1943) found that the Peking
variety of soybeans was killed at 0.4 mg/1, whereas the Manchu variety
was killed at 1.6 mg/1 zinc.
Although few data are available on the effects of zinc in the marine
environment, it is accumulated by some species, and marine animals contain
zinc in the range of 6 to 1,500 mg/kg (NTAC, 1968). As a goal, the
marine environment should be protected to the same level as the fresh
water environment.
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-------
REFERENCES CITED In TABLE 15:
(1) Brown, V.B., 1968. The calculation of the acute toxicity of mixtures
of poisons to rainbow trout. Water Res. 2:723.
(2) Brungs, W.A., 1969. Chronic toxicity of zinc to the fathead minnow,
Pimephales promelas Rafinesque. Trans. Am. Fish. Soc., 98:272.
(3) Cairns, J., Jr., e£ al_., 1971. The effects of pH, solubility, and
temperature upon the acute toxicity of zinc to the bluegill sunfish
(Lepomis macrochirus Rafinesque). Trans, of the Kansas Acad.
of Sci., 74(1).
(4) Chapman, G.A. Unpublished data available at the National Water Quality
Laboratory, Duluth, Minnesota.
(5) Garton, R.R., 1972. Biological effects of cooling tower blowdown.
71st National meeting. Amer. Inst. Chem. Engr. Jour., Dallas, Texas.
(6) Herbert, D.W.M., and D.S. Shurben, 1964. The toxicity to fish of
mixtures of poisons. I. Salts of ammonia and zinc. Ann. Appl.
Biol., 53:33.
(7) Hoi combe, G., and D.A. Benoit. Unpublished data available at the
National Environmental Research Laboratory, Duluth, Minnesota, 55C04.
(8) Mount, D.I., 1966. The effect of total hardness and pH on acute toxicity
of zinc to fish. Air/Water Poll., 10:49.
(9) Pickering, Q.H., and C. Henderson, 1966. The acute toxicity of some heavy
metals to different species of warm water fishes. Air/Water Poll.,
10:453.
-------
(10) Pickering, Q.H., and W.N. Vigor, 1965. The acute toxicity of zinc to
eggs and fry of the fathead minnow. Progressive Fish Culturist, 27:153.
(11) Rabe, F.W., and C.W. Sappington, 1970. The acute toxicity of zinc to
/
cutthroat trout Sa1mo C1 a r k i, pp 1-16. In: Biological Productivity
of the coeur d1 Alene River as related to water quality. Completion
Rept. Water Resources Res. Inst. Univ. Idaho, Moscow. Sport Fish.
Abs. 13665 1971.
(12) Rachlin, J.W., and A. Perlmutter, 1968. Response of an inbred strain
of platyfish and the fathead minnow to zinc. Progressive Fish
Culturist, 30:203.
(13) Rehwoldt, R., e_t a\_., 1971. Acute toxicity of copper, nickel, and zinc
ions to some Hudson River fish species. Bull. Environ. Contam.
- toxicol., 6:445.
(14) Rehwoldt, R., et _al_., 1972. The effect of increased temperature upon
the acute toxicity of some heavy metal ions. Bull. Environ. Contam.
Toxicol., 8:91.
(15) Sinley, J.R., et al_., 1974. The effects of zinc on rainbow trout
(Salmo qairdneri) in hard and soft water. Bull. Environ. Contam.
Toxicol., 12 No. 2.
(16) Spehar, R.L., 1974. Cadmium and zinc toxicity to Jordanella floridae
(Goode and Bean): Effects on growth, reproduction, survival, and
behavior. A thesis submitted to the University of Minnesota.
(17) Sprague, J.B., 1964. Lethal concentrations of copper and zinc for young
Atlantic salmon. Jour. Fish Res. Bd. Can., 21:17.
-------
(18) Sreenivasan, A., and R.S. Raj, 1963. Toxicity of zinc to fish.
Current Sci. (India), 32(8):363; Chem. Abs., 59:143386 (1963).
-------
REFERENCES CITED;
Affleck, R.J., 1952. Z1nc poisoning 1n a trout hatchery. Australian J ur
Marine Freshwater Res. 3:142.
Anderson, E.A.,et al_., 1934. . The corrosion of zinc 1n various
waters. Jour* Amer. Water Works Assn., 26:49.
Ball, I.R., 1967. The relative susceptibilities of some species of fresh
fish to poisons. I. Ammonia. Water Res. 1:767.
Bartow, E. and O.M. Welgle, 1932. Z1nc In water supplies. Indus. Eng.
Chem. 24:463.
Bleslnger, K.E. and G.M. Christensen, 1972. Effects of various metals on
survival, growth, reproduction, and metabolism of Daphnla magna. Jour,
F1sh. Res. Bd. Canada 29:1691.
Brungs, W.A., 1969. Chronic toxlclty of zinc to the fathead minnow, Plmephales
promelas Rafinesque. Trans. Amer. Fish. Soc. 98:272.
Cairns, j.t jr. and A. Scheler, 1958. The effect of periodic low oxygen upon
the toxlclty of various chemicals to aquatic organisms. 12 Ind. Waste Conf.
Proc. Purdue Univ. Eng. Extension Series.
Cohen, J.M., et.al_., I960. Taste threshold concentrations of
metals in drinking water. Jour. Amer. Water Works Assn. 52:660.
Early, E.B., 1943. Minor element studies with soybeans: 1. Varietal reactions
to concentrations of zinc 1n excess of nutritional requirements. Jour. Amer.
Soc. Agron. 35:1012.
-------
Herbert, D.W.M. and D.S. Shurben, 1964. The toxicity to fish of mixtures
of poisons. I. Salts of Ammonia and zinc. Ann. Appl. Biol. 53:33.
Hewitt, E.J., 1948. Relation of manganese and some other metals to the iron
status of plants. Nature, 161:489.
Hunter, J.G. and 0. Vergnano, 1953. Trace element toxicities in oat plants.
Ann. Appl. Biol. 40:761.
Kopp, J.F. and R.C. Kroner, 1967. Trace metals in waters of the United States.
U.S. Oept. Interior, Federal Water Pollution Control Adm., Cincinnati,
Ohio.
Lloyd, R., 1961. Effett of dissolved oxygen concentrations on the toxicity of
several poisons to rainbow trout (Salmo gairdnerii Richardson). Jour.
Exp. Biology 38:447.
Mlllikan, C.R., 1947. Effect of molybdenum on the severity of toxicity symptoms
in flax induced by an excess of either manganese, zinc, copper,, nickel, or
bobalt in the nutrient solution. Jour. Aust. Inst. Agr. Sci., 13:180.
National Technical Advisory Committee, 1968r Water Quality Criteria. Federal
Water Pollution Control Administration, Washington, D.C.
Pickering, Q.H., 1968. Some effects of dissolved oxygen concentrations upon
the toxicity of zinc to bluegill Lepomis macrochirus Rafinesque. Water
Res. 2:187.
Pickering, Q.H. and C. Henderson. 1966. The acute toxicity of some heavy metals
to different species of warm water fishes. International Jour. Air and Water
Poll., 10:453.
Skldmore, J.F., 1964. Toxicity of zinc compounds to aquatic animals, with
special reference to fish. Quarterly Rev. of Biology. 39:227,
-------
Sprague, J.B. and B.A. Ramsay, 1965. Lethal levels of mixed copper-zinc
solutions for juvenile salmon. Jour. Fish. Res. Bd. Canada. 22:425.
Vallee, B.L., 1957. 21nc and Its biological significance. Arch. Indust. Health,
16:147.
Warnick, S.L. and H.L. Bell, 1969. The acute toxIcHy of some heavy metals to
different species of aquatic Insects. Jour. Water Poll. Control Fed.
41:(Part 1)280.
Wurtz, C.B., 1962. Zinc effects on freshwater mollusks. Nautilus, 76:53.
-------
GLOSSARY
Acutely toxic: Causing death or severe damage to an organism
by poisoning during a brief exposure period,
normally ninety-six hours or less, although there
is no clear line of demarcation between acute and
chronic toxicity.
Chronically toxic: Causing death or damage to an organism by
poisoning during prolonged exposure, which,
depending 6n the organism tested and the test conditions
and purposes, may range from several days, to weeks,
months, or years.
Dose equivalent: The product of the absorbed dose from ionizing
radiation and such factors to account for differences in
biological effectiveness due to the type of radiation ajid
its distribution in the body as specified by the International
Commission on Radiological Units and Measurements
-------
Gross beta particle activity: The total radioactivity due to beta particle
emission as inferred from measurements on a dry sample
exclusive of the contribution, if any, due to potassium-40 and
other naturally occurring radionuelides.
LC25: The concentration of a toxicant that is lethal (fatal) to twenty-five
percent of the organisms tested under the test conditions in a
specified time.
LC50: The concentration of a toxicant which is lethal (fatal) to
fifty percent of the organisms tested under the test conditions
in a specified time.
LD50: The dose of a toxicant that is lethal (fatal) to fifty percent
of the organisms tested under the test conditions in a
specified time. A dose is the quantity actually administered
to the organism and is not identical with a concentration,
which is the amount of toxicant in a unit of test medium
rather than the amount ingested by or administered to the organism.
Liter (1): The volume occupied by one kilogram of water at a pressure
of 760 mm of mercury and a temperature of 4 C. A liter is
0. 9463 quart.
Man-made beta particle and photon emitters: All radionuclides emitting
beta particles and/or photons listed in Maximum Permissible
Body Burdens and Maximum Permissible Concentrations of Radio-
nuclides in Air or Water for Occupational Exposure, NBS Handbook
69, except the daughter products of thorium-232, uranium-235, and
uranium-238.
V9?
-------
Microgram per liter (ug/1): The concentration at which one millionth
of a gram (10~6g) is contained in a volurre of one liter.
There are 453. 59 grams in a pound.
Microgram per kilogram (ug/kg):- The concentration at which one millionth
of a gram (one microgram) is contained in a mass of one kilogram.
A kilogram is 2. 2046 pounds.
Milligram per kilogram (mg/kg): The concentration at which one thousandth
of a gram (one milligram) is contained in a mass of one kilogram. A
gram contains 1000 milligrams.
Milligram per liter (mg/1): The concentration at which one milligram (10~^g)
is contained in a volume of one liter.
Milliliter (ml): A volume equal to one thousandth of a liter.
Most Probable Number (MPN): The statistically determined number which
represents the number of individuals most likely present in a given
sample or aliquot, based pn test data.
Nanogram per liter (ng/1): The concentration at which one billionth
of a gram (10~ag) is contained in a volume of one liter..
5*00
-------
Part per million (ppm): A concentration at which one unit is
contained in a total of a million units. Any units may be
used (e.g., weight, volume) but in any given application
identical units should be used (e.g., grams per million grams
or liters per million liters).
Parts per thousand (o /oo): A concentration at which one unit is
contained in a total of a thousand units. The rules for
using this term are the same as those for parts per million.
Normally, this term is used to specify the salinity of estuarine
or sea waters.
Picocurie (pCi): That quantity of radioactive material producing 2. 22
nuclear transformations per minute.
Rem: The unit of dose equivalent from ionizing radiation to the total
body or any internal organ or organ system. A millirem (mrem)
is 1/1000 of a rem (0. 001 rem).
TL50: - Median Tolerance Limit: -The concentration of a test material
at which just fifty percent of the test animals are able to
survive under test conditions for a specified period of exposure.
Tim : Synonymous with
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